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Review of literature to determine the uses for ozone in the treatment of water and wastewater
Final Report
11/02/2013
Published by CREW – Scotland’s Centre of Expertise for Waters. CREW connects research and policy,
delivering objective and robust research and professional opinion to support the development and
implementation of water policy in Scotland. CREW is a partnership between the James Hutton
Institute and all Scottish Higher Education Institutes funded by the Scottish Government.
This document was produced by:
Lisa Avery
The James Hutton Institute
Craigiebuckler
Aberdeen
Peter Jarvis & Jitka Macadam
Cranfield Water Science Institute
Cranfield University.
Please reference this report as follows: Avery, Jarvis and Macadam (2013) Review of literature to
determine the uses for ozone in the treatment of water and wastewater.
Dissemination status: Unrestricted
All rights reserved. No part of this publication may be reproduced, modified or stored in a retrieval
system without the prior written permission of CREW management. While every effort is made to
ensure that the information given here is accurate, no legal responsibility is accepted for any errors,
omissions or misleading statements. All statements, views and opinions expressed in this paper are
attributable to the author(s) who contribute to the activities of CREW and do not necessarily
represent those of the host institutions or funders.
Contents
EXECUTIVE SUMMARY .......................................................................................................................1
1.0 DISINFECTION OF POTABLE WATER USING OZONE ...................................................................8
1.1 INTRODUCTION TO OZONE ................................................................................................................... 8
1.2 DISINFECTION CAPABILITY OF OZONE ................................................................................................... 11
1.3 OZONE INACTIVATION OF CRYPTOSPORIDIUM ...................................................................................... 13
1.3.1 Variability across studies on ozone inactivation of Cryptosporidium spp. .............................. 15
1.3.2 Inactivation of Cryptosporidium compared with other microorganisms ................................ 16
1.4 FACTORS AFFECTING OZONE DISINFECTION ........................................................................................... 17
1.5 OZONE DOSING AND CONDITIONS ....................................................................................................... 21
1.6 DISINFECTION BY-PRODUCTS ............................................................................................................. 23
1.6.1 Types and concentrations of DBP formed with ozonation ...................................................... 25
2.0 OXIDATION OF CONTAMINANTS IN WATER ........................................................................... 33
2.1 DESTRUCTION OF ORGANIC AND INORGANIC COMPOUNDS USING OZONE .................................................. 33
2.1.1 Iron and manganese removal ................................................................................................. 34
2.1.2 Oxidation of organic macropollutants .................................................................................... 35
2.1.3 Removal of micropollutants .................................................................................................... 36
2.1.4 Taste and odour control .......................................................................................................... 37
2.1.5 Particle removal ...................................................................................................................... 37
3.0 POINT OF USE DISINFECTION OF WATER SUPPLIES ................................................................. 38
3.1 APPLICATIONS AND DOSING CONDITIONS ............................................................................................. 38
4.0 APPLICATION OF OZONE IN WASTEWATER TREATMENT ........................................................ 39
4.1 DISINFECTION OF FINAL EFFLUENT ...................................................................................................... 40
4.2 IMPACT OF OZONE ON FILAMENTOUS BACTERIA .................................................................................... 42
4.3 OXIDATION OF INORGANIC COMPOUNDS ............................................................................................. 47
4.4 OXIDATION OF ORGANIC COMPOUNDS ................................................................................................ 47
4.5 ENHANCEMENT OF SLUDGE DEGRADATION ........................................................................................... 51
5.0 ENVIRONMENTAL IMPLICATIONS .......................................................................................... 51
6.0 CONCLUSIONS & RECOMMENDATIONS ................................................................................. 52
7.0 FUTURE RESEARCH AND DIRECTION ...................................................................................... 53
8.0 REFERENCES ......................................................................................................................... 54
List of Tables
Table 1. Water Quality based determination of ozone point of application for drinking water
treatment – (Modified from DeMers and Renner 1992; USEPA 1999). ....................................................... 9
Table 2. Effects of Ozone addition on common water treatment processes. ............................................ 10
Table 3. Overview of ozone applications in water treatment (adapted from Langlais et al., 1991). ......... 12
Table 4. Relative disinfecting capacity of ozone against a range of microorganisms. Data gives CT
(mg.min/L) values for 99% inactivation at 5 °C. Adapted from Langlais et al. (1990). Data in bold
indicates the best performing disinfectant. ................................................................................................ 13
Table 5. Cryptosporidium oocysts detected in 40-L ozonated and non-ozonated treatment trains
(after Nieminski and Brandford 1990). ....................................................................................................... 14
Table 6. Inactivation Kinetics of Cryptosporidium and other pathogens or indicator organisms at pH 7
(adapted from Von Gunten et al, 2003)...................................................................................................... 16
Table 7. Inactivation of a range of microorganisms in natural river water from Ohio, based on simple
CT values (after Owens et al., 2000). .......................................................................................................... 17
Table 8. CT values (mg.min/L) for inactivation of Cryptosporidium oocysts by ozone (after IEPA 2011). . 18
Table 9. Advantages and disadvantages of ozone mixings systems. Adapted from USEPA (1999). .......... 22
Table 10. Ozonation case study sites for potable water disinfection. ........................................................ 23
Table 11. Some key DBPs generated by Ozone and their Health Effects (adapted from Lipmann,
2009). .......................................................................................................................................................... 26
Table 12. DBP Regulations worldwide (adapted from Hrudley and Charrois, 2012).................................. 27
Table 13. Bromate formation at full scale WTWs in Europe and N America. Adapted from von Gunten
(2003b). ....................................................................................................................................................... 28
Table 14. Regression coefficient value for bromate formation from various locations. Adapted from
Jarvis et al. (2007). ...................................................................................................................................... 31
Table 15. Removal of iron and manganese in various case studies (adapted from Langlais et al.,
1991). .......................................................................................................................................................... 35
Table 16. Ozone demand for oxidation of organic compounds. ................................................................ 35
Table 17. Degree of removal of trace organics during ozonation in full-scale drinking water treatment
plants (Gottschalk et al., 2000). .................................................................................................................. 36
Table 18. Filamentous bacteria found in bulking activated sludge (adapted from Guo and Zhang,
2012). .......................................................................................................................................................... 43
Table 19. Ozone applications in wastewater treatment............................................................................. 49
List of Figures
Figure 1. Usual range of dosing positions for ozonation in bulk disinfection of drinking water. ............... 10
Figure 2. Illustrating the temperature dependence of CT values for 2-log inactivation of C. parvum in
natural waters at 3 °C (open triangles), 7 °C (filled triangles), 10 °C (open squares), 20°C (filled
squares), 22 °C (open circles) and 25 °C (filed circles). Data source: Oppenheimer et al., (2000). ............ 19
Figure 3. Schematic diagram of the reaction of organic and inorganic DBP precursors with
disinfectants to form regulated and emerging DBPs (After Krasner, 2009). .............................................. 24
Figure 4. Floc density with and without treatment (from Wijnbladh, 2007). ............................................ 44
Figure 5. Filament presence with and without treatment (from Wijnbladh, 2007). .................................. 45
Page | 1
Executive Summary
Disinfection of potable water using ozone
In most water treatment plants ozone is used for multiple applications. Ozone is now used as a
disinfectant, an oxidant of organic and inorganic molecules, a coagulant aid, removing taste and
odour, a means of controlling algae and as a way of biologically stabilising water. Ozone is very
effective for disinfection against bacteria, viruses and protozoa. However, when used in a
disinfection capacity, it is often used when contaminants are highly resistant to more conventional
disinfectants.
It has been described as the only chemical form of disinfection to provide effective inactivation of
Cryptosporidium and Giardia at doses similar to those used routinely for water treatment. For
drinking water, the CT concept or derivations of it are usually suitable for determining the required
dosage. Inactivation of Cryptosporidia in final effluent following wastewater treatment differs from
drinking water treatment due to the different water quality parameters. This renders the CT
approach much less effective for wastewater disinfection.
There is substantial variability across studies on ozone inactivation of Cryptosporidium spp. which
makes them difficult to summarise or directly compare. This is due to differences in the way studies
are performed (e.g. lab vs. pilot vs. full scale; synthetic or real water sources, artificial seeding with
oocysts and continuous vs. discontinuous ozone supply) and way in which the values are reported,
for example applied or transferred/residual ozone doses; differences in the level of detail with
respect to water quality parameters.
Despite this, log inactivation of Cryptosporidium following ozone application is frequently reported
to be within the 2-3 log range. Of the literature data evaluated graphically (45 data points from 12
studies), 61 % showed greater than 2-log inactivation. The CT relationship was not clear due to the
study differences highlighted above. Because Cryptosporidium is highly resistant to external
stressors including disinfectants, log inactivation cannot be reliably derived from studies of other
pathogens or indicator organisms. For example, in a study evaluating Cryptosporidium inactivation in
river water, Owens et al (2000) demonstrated that the CT required for 2-log inactivation of
Cryptosporidium (C. parvum and C. muris) was 12 times greater than that required for the same
This paper responds to a CREW call down request submitted by Scottish Water.
A review of literature to determine the uses for ozone in the treatment of water and wastewater; in particular:
point of use disinfection of water supplies
inactivation of cryptosprodium oocysts
bulk disinfection of potable water supplies at treatment works
oxidation of organics, iron and manganese
other uses
Wastewater:
impact of ozone on filamentous bacteria
disinfection of final effluent
other uses
Page | 2
degree of inactivation of Giardia muris cysts. Bacillus subtilis spores show some promise as
conservative indicators for Cryptosporidium inactivation, however.
Sequential disinfection schemes involving two disinfectants have been proposed to provide a certain
level of synergism, which may allow greater levels of inactivation to be achieved in drinking water
treatment plants (Gyurek et al., 1996; 1997; Liyanage et al., 1997). Rennecker et al. (1999b)
indicated that sequential disinfection with ozone followed by free chlorine was promising for the
treatment of oocysts.
A number of water quality parameters strongly influence the efficacy of the ozone to disinfect:
Temperature: as temperature increases, the disinfecting power of ozone increases.
Temperature influences inactivation of Cryptosporidium, with a 4.5-fold increase in CT
suggested for a 10 degree C rise in water temperature.
pH: changes in water pH changes the balance of available O3 and OH. An increase in pH from
6 to 9 reduces the amount of O3 available for disinfection by a factor of 40 (Elovitz et al.,
1999). There is conflicting evidence in the literature relating to the effect of pH on
inactivation of Cryptosporidium. It has been suggested that in batch systems, increasing pH is
correlated with increasing inactivation, but where ozone is continuously bubbled through
the system there is limited effect of pH. Inactivation rates of different Cryptosporidium
species were not substantially different, although a few studies have noted that oocysts of
different ages may show differential responses to ozonation.
Suspended solids and other ozone scavengers: the presence of contaminants other than the
target microorganisms may consume ozone and reduce the disinfection capacity of the
water. As for other disinfectants, constituents within the water, primarily NOM, EfOM,
BDOC, bromide, synthetic organic compounds and alkalinity exert an ozone demand. It is
therefore critical to know what is in the water to understand ozone doses and contact
necessary to achieve a specific water quality objective.
When applying ozone at a WTW facility, there are four requisite components: (1) oxygen gas feed
system (either air or pure oxygen); (2) ozone production and delivery, (3) an ozone contactor and (4)
ozone off gas destruction.
One of the key challenges faced when using ozone as an oxidising agent is the formation of
disinfection by-product (DBPs) compounds. Ozone tends to produce DBPs in the categories of
oxyhalides, aldehydes and carboxylic acids. While many organic and inorganic ozonation
disinfection/ oxidation by-products have been identified, bromate is generally considered to be of
greatest concern (von Gunten, 2003) and aldehydes are also important although they are not
currently regulated (Silva et al., 2010). Where bromide is present in raw waters, bromo-organic by-
products can form during ozonation.
There are a number of factors that influence bromate formation. These are:
Bromine concentration: given that bromide is oxidised by ozone to bromate, an increase in
bromide leads to an increase in bromate for a constant ozone dose and contact time.
pH: As the pH of the water is increased during ozonation, more bromate is formed.
Alkalinity: The presence of inorganic carbon (IC) species increases bromate formation.
Ammonia concentration: Ammonia can remove a significant intermediary from the bromate
formation path and reduce the amount of bromate formed.
Page | 3
Transferred ozone dose and contact time: The relationship between bromate formation and
CT follows a linear function, with an increase in CT leading to an increase in bromate
formation
The formation of disinfection by products of ozonation has been studied with respect to water
reclamation and is also pertinent where treated effluent significantly influences raw water for
abstraction.
In a study evaluating DBP concentrations across 12 drinking water treatment plants in the US in
which sites using all four major disinfectants (chlorine, chloramines, ozone, and chlorine dioxide)
were covered, the highest concentration of the DBP dichloroacetaldehyde occurred at a plant using
chloramine and ozone disinfection. Therefore, although the use of alternative disinfectants
minimized the formation of the four regulated THM, some unregulated DBPs were present in higher
levels than where traditional chlorine disinfection was applied. The literature is contradictory, and
ozonation can increase THMS where organic loadings of wastewater are still high, however it has
also been shown to reduce the formation of both THMs and HAAs where ozone is introduced in
combination with chlorination of effluent.
Oxidation of contaminants in water
Compounds present in water can react with ozone directly or indirectly through OH radicals. Direct
ozonation is usually the most important oxidative reaction if the radical reactions are inhibited due
to the lack of initiating compounds to begin the chain reaction or due to the presence of too many
radical scavengers. The direct pathway normally dominates under acidic conditions (pH <4) and
changes to the indirect pathways above pH 10. Both pathways will therefore play a role in most
ground and surface waters (pH ~ 7).
A number of compounds can be directly degraded by ozone. These include taste and odour
compounds (geosmin and methylisoborneol (MIB)), phenolic compounds and pesticides such as
atrazine. The importance of ozonation in the treatment of industrial wastewaters targeting the
degradation of dyes, pharmaceuticals and personal care products has also grown in recent years.
Ozone is also used for oxidation of organic macropollutants and its application is used for bleaching
of colour, increasing the biodegradability of organic compounds, removal of THM precursors and
reducing total organic halide formation potential or chlorine demand. One of the most important
ozone applications in water treatment is the oxidation of iron and manganese.
Ozone is capable of destroying a range of volatile micropollutant compounds, in particular alkenes
and aromatic organics under the conditions of treatment applied to drinking water. In the past,
micropollutant removal was not a primary task for ozone but was considered a positive side effect.
However, due to ever lowering detection limits and stricter regulatory requirements for more
chemicals in drinking water, the interest in micropollutants has grown in recent years.
The most common taste and odour associated compounds are MIB and geosmin and these have a
very low reactivity with ozone. However, despite this, studies with natural waters have shown good
removal efficiencies of these compounds when using ozone. It is likely that ozonation is most
effective in waters that support the OH radical pathway. However, the action of ozone in natural
waters is variable and depends on the quality of organics present as well as the treatment
conditions.
Page | 4
It has been observed for more than 30 years that pre-ozonation ahead of solid-liquid separation
processes can improve the removal of particles. Positive effects of pre-ozonation are also seen for
algae removal. Ozone readily kills or lyses many types of algae and it has also been observed to
enhance the removal of algae by coagulation and settling. Ozone can also be applied to inactivate
zooplankton and actinomycetes. A number of laboratory studies have reported the effect of
ozonation on the removal of cyanotoxins and it was shown that complete removal of the toxins can
be achieved when ozone is included in the treatment process. Ozoflotation is a new process
combining the physical phenomenon of flotation with the oxidising properties of ozone and is
usually used as a pre-treatment stage in order to reduce the treatment load.
Point of Use systems
There are a range of commercially available point of use (POU) and point of entry (POE) ozonation
devices. There is limited scientific literature available on the performance and reliability of these
devices and most of the below information has been taken from commercial sources and, as such,
limited validation of performance can be gleaned from this data. Independent testing of POU ozone
devices is required because to date, most, if not all, claims made by manufacturers have not been
verified.
Application of ozone in wastewater treatment
Similarly to drinking water treatment, ozone can be applied to satisfy a number of objectives in
wastewater treatment, including:
Disinfection
Oxidation of inorganic compounds
Oxidation of organic compounds
Enhancement of sludge degradability
Ozone disinfection mechanisms in wastewater are less well understood. Ozone reacts strongly with
many substances, therefore it is generally deemed more appropriate for use on pre-treated effluents
(Paraskeva and Graham, 2002). In a wastewater of high organic content, inactivation of total and
faecal coliforms and E. coli has been reported at an applied ozone dose of 10 mg/L for a 5 minute
contact time. Ozone doses commonly presented in the literature ranged from 0.3 µg/L to fully
saturated, with contact times generally between 1.5 and 18 minutes. This provided a range of
inactivation rates, broadly in the range of 1-2.5 log inactivation for coliforms, Enterococci and
Clostridia, while some higher reductions were noted for bacteriophages used as surrogates for
human viruses
The critical parameter for disinfection of indicator organisms appears to be the optimisation of mass
transfer of ozone, which is usually low due to its poor solubility. Hydraulic retention time appears
less important. Meeting the initial ozone demand leads to a substantial microbial inactivation as the
microbial cells actually exert a significant proportion of that ozone demand (Xu et al 2002).
Filamentous organisms are a normal part of the activated sludge microflora and, in low numbers, are
thought to promote floc formation. Excessively long filaments or presence in high numbers can lead
to sludge bulking (Eckenfelder, 1992). Ozone can be used as a “non-specific” approach to reducing
filamentous organisms and therefore reducing bulking and foaming in activated sludge systems. Low
Page | 5
dose ozonation can inhibit the activity of filamentous bacteria and has been applied to control
bulking and improve floc settling (Foladori et al, 2010).
The efficiency of sludge ozonation depends on the following parameters (Foladori et al; 2010):
Wastewater or Sludge quality
Reactor configuration
Ozone gas flow rate and concentration
Sludge flow rate and solids concentration
Ozone transfer efficiency
Contact time
Ozone dosage per mass of TSS
Studies have reported improved floc structure directly after the start-up of ozone treatment with
few filamentous bacteria remaining inside the sludge flocs. One author reported the number of
filaments to be an order of magnitude lower in the ozonated treatment than the control. The
authors also indicted that ozonation promoted nitrification and biological removal of organic
material without affecting phosphate removal.
The influence of ozone on other wastewater treatment parameters is critical to the success of its use
to reduce filaments or excess sludge. Paul and Debellefontaine (2007) noted that there was a linear
relationship between the log of biomass activity (reported as maximal oxygen uptake rate) and log
ozone dose between 0.001 and 0.2 g O3 transferred per g COD in the sludge). They concluded that
ozonation does not affect any of the capabilities of an AS biological process; however other studies
have reported contrary indications. For example, a decrease in nitrification rate has been shown to
be proportional to increasing ozone dose (Dyctzak et al (2007), although other studies indicate no
effects on nitrification. Reid et al (2007) suggested that a more cost effective ozone process for
excess sludge reduction would follow the principle “partial oxidation as low as possible and
biological oxidation as high as possible”. This minimises the use of ozone but maximises conversion
of solids into biodegradable materials which can then be removed in cheaper biological reactors.
Water quality and ozone demand can be used to determine the best point of application. Broadly,
high ozone demand in raw water, indicative of high levels of organic material can lead to increased
biodegradability of natural organic matter (NOM) (Beltrán et al. 1999; Ternes et al. 2006) following
ozonation. This may then require a biological treatment step to remove BDOC which can lead to
bacterial growth in the distribution system. Water with a high ozone demand and high turbidity
would be considered the most difficult water to treat with ozone, for example, surface water with a
high loading of organic material and inorganic particles therefore selecting a point in the treatment
train where particulates and organic matter have been removed substantially would be advisable.
However, because ozone can also break down organic material, antibiotics and pharmaceuticals,
incorporating ozonation at more than one stage of treatment may be effective. Thus low doses of
ozone prior to existing treatment can capitalise on some of these effects.
For wastewater, ozonation is often part of a multistage treatment process to reduce ozone
consumption. Most often a chemical-biological process includes biodegradation at least before and
also often after the chemical oxidation step (Gottschalk et al, 2010). Ozone quantities and
generation costs tend to be reduced when utilising these combined systems.
Page | 6
Dissolved organic carbon (DOC) concentrations are typically greater in wastewater than in surface
water, resulting in faster O3 decomposition rates and increased hydroxyl radical ( OH) exposures
(Buffle et al., 2006). As a result, higher O3 dosages are required to meet wastewater treatment goals,
potentially leading to increased DBP formation.
One of the aims of ozone application in wastewater treatment is to remove toxic inorganic
substances and this mainly involves the removal of cyanide (CN-) mostly associated with metal
processing and electronics industry wastewaters. Nitrite (NO2-) and sulphide (H2S/S2-) react quickly
with ozone and therefore their removal is sometimes carried out using ozonation, however more
cost-efficient biological treatment alternatives are more often employed for these contaminants.
Most often in industrial wastewater treatment, ozone is applied to remove target organic
compounds that can be present at a wide range of concentrations. These wastewaters include
landfill leachate, textile, pharmaceutical and chemical industry wastewaters that can contain many
refractory organics including humic compounds, aromatic compounds containing metals, pesticides
and surfactants. The main aims of ozonation in this case are (Gottschalk et al., 2000):
The transformation of toxic organics that are often present in low concentrations and as
complex mixtures
The improvement of biodegradability of refractory organics by partial oxidation
The removal of colour
Ozonation is a widely used chemical method to improve anaerobic degradability of sludge.
Ozonation has also been combined with anaerobic digestion as a pre-treatment or post-treatment
with a recycle back to the anaerobic digester. Better performance and lower ozone consumption has
been observed in the case of post-treatment and recycling in the digester. The advantages of
ozonation pre-treatment of secondary sludge is that it can improve the sludge solubilisation and it
can also simultaneously degrade organic pollutants.
Conclusions and recommendations
Key consideration of ozone treatment for any of the above processes include improved
ozone transfer technologies, correct dosage and an awareness of the importance of key
water quality parameters such as organic, particulate and bromide loadings.
Implementation of ozone treatments is likely to be most effective if evaluated on a site by
site basis as this not only promotes a thorough evaluation of how to implement the
technology for effective disinfection, but would also allow the operators to determine
whether ozone interventions at multiple points would be more cost effective for overall
treatment and/or whether the inclusion of additional treatment stages before or after
ozonation would be required.
Ozone disinfection of drinking water is highly effective against a wide range of pathogens,
including Cryptosporidium.
Cyst and sporular forms of protozoans and bacteria present the most difficult treatment
challenge for ozone disinfection. High CT values can control these pathogens, but it is
recommended that a physical barrier is also provided for water sources containing these
micro-organisms.
Page | 7
The main limitation of applying ozone in drinking water systems is the production of
bromate as a DBP. Methods of controlling bromate formation have been developed,
principally through controlling pH and understanding prevailing water quality conditions.
POU/POE systems are widely available, mainly from North American suppliers that can
generate and deliver ozone from mains electricity. A filtration system needs to be supplied
after the ozone when these are used in order to prevent precipitated solids from being
present in treated water – it is not always evident that the proprietary systems have these as
supplied. Who manages and replaces spent filtration systems and adequate off-gas control
must therefore be also considered for POU systems.
A wide range of organic and inorganic contaminants can be degraded by ozone. High
concentrations of ozone are needed for effective degradation of bulk natural organic matter
and is therefore not recommended for this application. The most effective use of ozone is
for oxidation of metals and organic micro pollutants in combination with a
physical/adsorptive process.
The higher contaminant and scavenging load in wastewater means that much higher doses
must be applied for these waters in order to achieve satisfactory levels of
removal/disinfection. There are many examples of ozone having been used in wastewater
for small scale and pilot treatment systems, but few cases of large WWTWs using ozone due
to the high cost of having to add such high concentrations of ozone.
Ozonation of final effluent can be effective, however higher ozone doses tend to be required
which increases the likelihood of formation of disinfection by products.
Ozone can be an effective non-specific inhibitor of filamentous organisms in activated sludge
systems but effects on the overall treatment process are variable
Ozone is a widely used chemical in water and wastewater treatment as well as numerous
industrial applications. The ability of ozone to effectively oxidise a wide range of
contaminants and disinfect a broad sweep of micro-organisms have made it an essential
component of many treatment flowsheets across the world.
Key words
Disinfection, iron, manganese, ozone, oxidation, point of use, wastewater, water.
Page | 8
1.0 Disinfection of potable water using ozone
1.1 Introduction to ozone
In most water treatment plants ozone is used for multiple applications: as a biocide (disinfection), as
an oxidant or as a pre-treatment to improve performance of subsequent processes (for example,
prior to GAC, sand filtration and coagulation) (Langlais et al., 1991). Due to these multiple uses,
ozone may be applied at a number of points in the water treatment train (Figure 1). At all of these
stages, ozone may disinfect the water, but its primary use may be for another application (for
example oxidation or enhancing biodegradability). At many locations, multiple stages of application
of ozone (usually two) have proven to be the most appropriate and cost effective way of meeting
specific water quality objectives. The action of ozone may be enhanced by combining ozone with
additional chemicals or physical processes to enhance its mode of action. This includes the
downstream of adsorption processes such as GAC or the addition of hydrogen peroxide, UV
radiation, ultrasound and metallic catalysts (such as reduced iron). In these cases, the aim is to
promote the formation of reactive radicals that may enhance the degradation of contaminants and
microbes (Glaze, 1987).
For drinking water treatment, ozone is typically applied to the raw water or the water after
flocculation and clarification. Water quality and ozone demand can be used to determine the best
point of application (Table 1). Broadly, high ozone demand in raw water, indicative of high levels of
organic material can lead to increased biodegradability of natural organic matter (NOM) (Beltrán et
al. 1999). Here, large organic molecules are converted into smaller organic molecules of more
biodegradable dissolved organic carbon (BDOC). This may then require a biological treatment step to
remove BDOC which can lead to bacterial growth in the distribution system. High turbidity may
indicate the presence of particulate matter which may contain both organic and inorganic
components. Organic content is considered to exert more influence on the ozone demand than
suspended solids (Janex et al., 2000). The presence of oxidizable organic constituents or bromide
ions will generate disinfection by-products upon ozonation. Water with a high ozone demand and
This paper responds to a CREW call down request submitted by Scottish Water.
A review of literature to determine the uses for ozone in the treatment of water and wastewater; in particular:
point of use disinfection of water supplies
inactivation of cryptosprodium oocysts
bulk disinfection of potable water supplies at treatment works
oxidation of organics, iron and manganese
other uses
Wastewater:
impact of ozone on filamentous bacteria
disinfection of final effluent
other uses
Page | 9
high turbidity would be considered the most difficult water to treat with ozone, for example, surface
water with a high loading of organic material and inorganic particles (Table 2). Ozone generally leads
to a reduction in disinfectant demand of finished water (USEPA, 1999).
Ozone is frequently used before coagulation to help the downstream coagulation efficiency (Langlais
et al., 1991). However, this is usually for lowland reservoir water sources of low organic content.
Ozone has been reported to have widely varying effects on the adsorption of NOM to aluminium
hydroxide. In raw water, ozone appeared to be more reactive with the humic (FA and HA) fraction of
NOM. This led to a decrease in the efficacy of a subsequent alum coagulation step for NOM removal.
If humic substances were removed by pre-coagulation with alum and then ozonated, the remaining
fractions adsorbed more effectively to alum (Bose et al., 2007). The authors therefore recommended
a strategy of staged coagulation with intermediate ozonation for waters containing both humic and
non-humic NOM, in order to achieve maximal removal of dissolved organic carbon.
Ozone can also increase the biodegradability of other compounds such as antibiotics and other
pharmaceuticals including fibrates (which lower LDL cholesterol) and estrogens (Larcher 2012 and
references therein). Thus low doses of ozone prior to existing treatment can capitalise on some of
these effects.
Table 1. Water Quality based determination of ozone point of application for drinking water treatment – (Modified from DeMers and Renner 1992; USEPA 1999).
Ozone Demand
Turbidity Water Characteristics
Potential Issues
Where to add: Reason
Low Low High quality raw water
- Raw water Only practical point of application
Low High Inorganic materials (clay, silt)
May produce some DBPs
After pre sedimentation or conventional sedimentation
Reduced ozone demand
High Low Dissolved constituents e.g. bromide, iron, manganese, colour, organics
Production of DPB and/ or BDOC
Raw water or post sedimentation. If biodegradable organics produced, best upstream of biological treatment.
Biological treatment may be required to remove biodegradable organics
High High High concentrations of organic and inorganic materials
Production of DPB and/ or biodegradable organics
After sedimentation and possibly after filtration If v. high O3 demand, dual O3 feed points may be required
V. high organics may require biological treatment
Page | 10
Figure 1. Usual range of dosing positions for ozonation in bulk disinfection of drinking water.
Table 2. Effects of Ozone addition on common water treatment processes.
Process Effects References
Biologically Active Filtration (BAF)
↑biodegradability of dissolved organics
↑ DO2 and efficacy of biological filters
↓ BDOC
Removes NOM, reducing potential for DBP formation
↓ Demand for residual disinfectant
Helps remove by products of ozonation
USEPA (1999)
Slow Sand Filters O3 addition before slow sand filtration ↑TOC and BDOC removal
Rachwal et al. (1988); Zabel (1985); Eighmy et al. (1991); Malley et al. (1993)
Rapid Rate Filters May ↓ assimilable organic carbon (AOC) but extent of BDOC unclear
USEPA (1999)
All filtration O3 oxidation of iron and manganese ↑insoluble oxides and requirement for↑sedimentation or filtration. ↑ backwash frequency.
USEPA (1999)
Granular Activated Carbon
↑ efficiency of GAC removal of BDOC
Variable removal of AOC.
Katz (1980); Langlais et al. (1991)
Disinfection Usually ↓ subsequent chlorine, chlorine dioxide, or monochloramine demand of finished water
Possible ↓ NH4 + and ↑in NO3
- presence of
ammonium (literature contradictory)
Martinez et al. (2011) and references therein
Raw water Filtration Coagulation /flocculation
Solid separation
Disinfection
Pre - ozonation Inter-ozonation Post-ozonation
Role: Disinfectant
Coagulant aid Oxidation
Role: Disinfectant Oxidation
Bio - stabilisation
Role: Disinfectant
Page | 11
1.2 Disinfection capability of ozone
Although the widespread use of ozone across the world did not occur until the 1950-60’s, the
understanding that ozone was an extremely effective way of disinfecting polluted water had been
apparent since the end of the 19th century (Langlais et al., 1990). Ozone is now used as a
disinfectant, an oxidant of organic and inorganic molecules, a coagulant aid, removing taste and
odour, a means of controlling algae and as a way of biologically stabilising water (Table 3). However,
in most full-scale applications, the main driver for implementing ozone is for disinfection purposes
(von Gunten, 2003a). This is because ozone is an exceptionally good disinfectant that has faster
disinfection kinetics and is more potent to most microorganisms when compared with other widely
used chemical disinfectants (Table 4). Ozone is very effective for disinfection of bacteria, viruses and
protozoa. However, when used in a disinfection capacity, ozone is often preferentially used when
contaminants are present that are highly resistant to more conventional disinfectants. For example,
as is expanded on elsewhere in this report, ozone is often chosen for disinfection of water
contaminated with protozoan cysts (such as Cryptosporidium and Giardia) because it is far more
effective at destroying these organisms at low concentrations and contact times when compared
with chlorine based chemicals.
The CT concept (CT being the disinfectant concentration multiplied by the contact time) is an
underpinning approach for determining and evaluating disinfection processes in drinking water
without attempting to measure actual pathogen inactivation where pathogen numbers are generally
too low to enumerate (Broséus et al., 2008). Disinfection is brought about through the combination
of disinfectant dose and the length of time for which a given organism is in contact with that
chemical. Thus, for inactivation to occur, either a high dose and short contact time or a lower dose
but a longer contact time is required. Typically used in the design of drinking water disinfection
systems, at its simplest, the CT value is obtained by multiplying the disinfectant residual
concentration, C, by the contact time, T, in the water system from the point where the disinfectant is
applied to the point where the residual is measured (USEPA, 2009). C is usually measured at the end
of the contactor, therefore reflecting minimal oxidant concentration and T is usually represented by
T10, the time taken for 10% of the water to pass through the reactor. The basic CT10 concept, while
conservative does not take into account the effects of water quality (other than temperature and pH
which are incorporated into CT reference tables) and modifications of the equation have been
developed (e.g. Broséus et al, 2008) with the aim of providing better estimates of CT under full scale
conditions. These are extensively reviewed by Rakness (2005).
The linear Chick-Watson model describes the inactivation of microorganisms with chemical
disinfectants as a first order chemical reaction.
ln (N/No) = -kCnt where:
N = number of microorganisms present at time t
N0 = number of microorganisms present at time 0
k = coefficient of specific lethality
C = coefficient of disinfectant
n = coefficient of dilution (constrained to unity in the linear form)
t = time
Page | 12
A constant rate of kill is not realistic for some organisms and disinfectants and models incorporating
additional parameters, for example to account for tailing off in survival curves (Hom model) have
been applied. However, they do not always provide a better fit (Oppenheimer et al., 2000) in which
case the linear Chick-Watson model is as effective.
Advanced CT models provide an effective tool for evaluating effective microbiological CT of
disinfection processes under different conditions and controlling CT can reduce DBP formation
(Broséus et al, 2008). The CT required to achieve a given level of inactivation is unique (Driedger et
al., 2000), although some authors have suggested that doses of 5-15 mg ozone per L is required for
adequate disinfection (Martinez et al, 2011). Since long contact times can only be facilitated by large
volume contacting systems, which have large footprints and require large capital investments, high
dissolved ozone concentrations are likely to be advantageous Meyer et al. (2000).
Table 3. Overview of ozone applications in water treatment (adapted from Langlais et al., 1991).
Application Point of ozonation in the treatment train
Ozone dose Best pathway
Comments
Iron and manganese removal
pre-, inter- medium direct inter- application may be best with high DOC waters
Colour removal inter- medium to high direct two-step stoichiometry
Taste and odour control
inter- high indirect T and O may be produced by low ozone doses
Synthetic organic compounds
inter- medium to high indirect direct pathway can be best for some compounds
Particles pre- low unknown may require high calcium concentration
Algae pre-, inter- Low to medium unknown can be used with flotation
Pathogens pre-, post- medium to high direct pre- (US), post- (Europe)
Chlorination disinfection by-products
pre-, inter- Low to high direct high levels of removal require radicals (indirect pathway)
Biodegradable compounds
inter- medium unknown design of downstream filtration process is important
Escherichia coli has been shown to be reduced by 4-logs in less than 1 minute at 0.09 mg/L ozone
residual concentration. Legionella pneumophilia was reduced by over 2-logs when contacted with an
ozone level of 0.21 mg/L for 5 minutes. Similar sensitivity to ozone has been seen for Staphylococcus
and Pseudomonas and Salmonella. The most resistant species to ozone were Streptococcus and
vegetative bacteria and those that produce sporular forms, such as Bacillus and Mycobacterium, but
all are effectively killed by relatively low ozone concentrations (US EPA, 1999). Viruses are also
effectively removed by ozone. Rotavirus, phage, polio and coxsackie viruses have all been
demonstrated to be removed by ozone. Over 5 log removal of coxsackie virus was observed at an
ozone dose for 1.45 mg/L in lake water giving a 0.28 mg/L ozone residual (US EPA, 1999).
Page | 13
Table 4. Relative disinfecting capacity of ozone against a range of microorganisms. Data gives CT (mg.min/L) values for 99% inactivation at 5 °C. Adapted from Langlais et al. (1990). Data in bold indicates the best performing disinfectant.
Micro-organism Ozone
pH 6-7
Chlorine
pH 6-7
Chloramine
pH 8-9
Chlorine dioxide
pH 6-7
E. Coli 0.02 0.034-0.05 95-180 0.4—0.75
Polio 1 0.1-0.2 1.1-2.5 770-3740 0.2-6.7
Rotavirus 0.006-0.06 0.01-0.05 3810-6480 0.2-2.1
Phage f2 - 0.08-0.18 - -
G. Lambia cysts 0.5-0.6 47->150 - -
G. Muris cysts 1.8-2.0 30-630 1400 7.2-18.5
1.3 Ozone Inactivation of Cryptosporidium
Cryptosporidiosis is a predominantly waterborne disease with infections caused by contaminated
drinking water, among other routes (HPA, 2008). Between 106 and 1011 Cryptosporidium oocysts
per g of human faeces are released into wastewater treatment plants and may be released into
surface waters if process failure occurs (Wohlsen et al., 2007). Furthermore, Cryptosporidium
oocysts are shed by livestock and are prevalent in surface waters from both human and animal
sources (Pintar et al., 2012). Therefore eliminating the organism during drinking and wastewater
treatment processes is important for the protection of human health. Cryptosporidium oocysts are
particularly resistant to disinfection (Campbell et al., 1982). Korich et al. (1990), in a study comparing
the efficacy of ozone, chlorine dioxide, chlorine and monochloramine against C. parvum, concluded
that, with the possible exception of ozone, disinfectants could not be expected to inactivate oocysts
in drinking water.
Ozone is a strong oxidant and has been shown to inactivate many different types of microorganism
(Appendices 1-4; Blanc et al.; 2005; John et al., 2005; Chand et al., 2007; Kobayashi et al., 2011). The
oxidizing mechanisms of ozone may involve direct reactions of molecular ozone and also free
radical-mediated destruction. It is thought to affect cellular components such as proteins lipids,
peptidoglycans, enzymes and virus capsids (Voidaru et al, 2007). Ran et al (2010) found that ozone
did not damage the DNA and RNA within the oocysts, but appeared to lead to the folding (at 60
seconds) and ultimately shrinking and bursting (at 8 minutes contact time; dose not given but
assumed to be 3 mg/L).
Due to the large CT values required for C. parvum inactivation, treatment with free chlorine may not
result in adequate oocyst inactivation under most conditions relevant to drinking water treatment.
Ozone has, however, been proved superior to chlorine and monochloramine for Cryptosporidium
parvum oocyst inactivation (Driedger et al., 2000; 2001) and has been described as the only
chemical form of disinfection to provide effective inactivation of Cryptosporidium and Giardia at
doses similar to those used routinely for water treatment (Irish EPA, 2011). Von Gunten (2003)
stated that ozone is “an excellent disinfectant and is able to inactivate even more resistant
pathogenic microorganisms such as protozoa (e.g. Cryptosporidium parvum oocysts) where
Page | 14
conventional disinfectants (chlorine, chlorine dioxide) fail”, although the author also noted that the
ozone exposure required to inactivate these microorganisms is quite high. Korich et al.(1990)
demonstrated a 99% (2-log) inactivation of oocysts with an ozone concentration of 1 mg/L for 5-10
minutes (Appendix 1), whereas to achieve the same results with chlorine required a dose of 80 mg/L
for 90 minutes. Nieminkski and Bradford (1990) applied ozone treatment to two of four parallel
continuous flow treatment trains in a pilot reactor employing coagulation/flocculation with ferric
chloride and direct filtration. Ozone was applied before coagulation at a CT value of 1.5 mg.min/L
(intended to provide 99.9% inactivation of Giardia lamblia cysts). Although not the principle focus of
the trials, Cryptosporidium oocysts were enumerated in the filter backwash water and were
detected only in the non-ozonated treatment trains (Table 5).
Table 5. Cryptosporidium oocysts detected in 40-L ozonated and non-ozonated treatment trains (after Nieminski and Brandford 1990).
Sampling occasion Raw Water Train 1 (O3) Train 2 Train 3 Train 4 (O3)
1 0 0 0 + 0
2 3 0 3 3 0
3 1 0 3 2 0
4 4 0 + + 0
Inactivation of Cryptosporidia in final effluent following wastewater treatment differs from drinking
water treatment because the different water quality parameters lead to a range of issues discussed
later in this review. However, several studies have reported on the efficacy of final effluent
disinfectant with ozone, specifically in relation to Cryptosporidium. In a quantitative microbial risk
assessment study, Pintar et al (2012) modelled the mean individual daily probability of
Cryptosporidium infection under existing conditions (advanced wastewater treatment with ozone
disinfection) and also under a scenario where ozone treatment was removed. They demonstrated
that the risk of infection increased, albeit less than ten-fold, if ozone treatment ceased. However, in
contrast, in a 6-year monitoring study of Cryptosporidium oocyst prevalence in source (river) water
in Seoul, including a sampling location heavily influenced by a sewage treatment works, Lee et al.
(2010) found no significant differences in numbers before and after the implementation of ozone
final effluent disinfection, indicating that it may be insufficient to inactivate Cryptosporidia. Liberti et
al. (2000) also reported that ozone was ineffective (Appendix 1) in inactivating Cryptosporidium
parvum, however their study utilised clarified secondary treated effluent and the same effluent post
filtration through sand and gravel. The initial loading of pathogens was low and it is likely, as the
authors concluded, that the low concentrations of C. parvum present (10 and 2 oocysts detected)
rendered the microbial counts inaccurate.
Overall, for (primarily) drinking water and final effluent, log inactivation of Cryptosporidium
following ozone application is frequently reported to be within the 2-3 log range (Appendix 1). Of
the literature data evaluated graphically (45 data points from 12 studies), 61 % showed greater than
2-log inactivation. The CT relationship was not clear due to the study differences highlighted above
(Figure 2).
Page | 15
Figure 2. Literature data on inactivation of Cryptosporidium spp. across a range of CT values. Data were derived from Appendix 1 utilising all studies for which suitable CT values and inactivation data were available. Open squares indicate consumed ozone was used in the calculation of CT; filled circles represent data where applied ozone was utilised.
1.3.1 Variability across studies on ozone inactivation of Cryptosporidium spp.
As with many microbial-related inactivation reviews, some difficulties arise when attempting to
summarise the literature relating to ozone disinfection of Cryptosporidium spp. Firstly, there are a
number of studies based on laboratory scale reactors (Appendix 1) and few which fully replicate full
scale treatment situations. The necessity of seeding test waters with oocysts in many cases is
evident from Liberti et al (2000). However, it is encouraging to note that Owens et al, using a pilot
scale reactor on natural river water from Ohio, observed comparable results to bench scale studies
(Owens et al, 2000). The reactor type is of particular importance as batch studies are most likely to
demonstrate proficient disinfection compared with flow through systems (Rochelle et al, 2005).
However, it is important to attempt to replicate the conditions of the proposed full scale process
when carrying out bench scale tests.
Secondly, data are reported in different forms throughout the literature and single parameters
cannot always be readily extracted, for example a CT value may be given or the specific
concentration and contact time details may be provided, with or without reference to whether the
authors refer to applied, residual or transferred ozone. In other cases, a rate constant for ozone
inactivation of Cryptosporidium is given. Furthermore, background parameters which may affect
ozonation reactions are not always provided in full and a number of studies utilised clean water or
buffer solutions only for disinfection trials (Korich et al, 1990; Driedger et al, 2001; Rennecker et al.,
1999; Ran et al, 2010a) which, while useful as a starting point do not accurately reflect full scale
treatment.
In addition to the above aspects, methods of determining inactivation can also differ significantly
among reports and studies. For example, the excystation approach is known to underestimate viable
0
10
20
30
40
50
60
CT
valu
e (
mg.
min
/ L)
Inactivation Rate Category (log reduction)
3 + >2 <3 1 <2
<1
Page | 16
oocysts and is likely to differ from animal infectivity approaches, which can be quite variable.
Burkhari et al (2000) illustrated variability of close to 99% in a comparison of different viability
determinations, including application of various stains (DAPI/PI, Syto 9, Syto 59) and excystation.
They also considered the relationship with infectivity using a mouse model. Determination of viable
and infective Cryptosporidium is discussed in depth by Rochelle et al. (2005). With this in mind,
Appendix 1 and Tables 1-3 provide a summary of some keys findings on ozone inactivation of
Cryptosporidium.
1.3.2 Inactivation of Cryptosporidium compared with other microorganisms
Owens et al (2000) tested the efficacy of ozone inactivation of a range of microorganisms in river
water from Ohio, using a pilot scale ozonation reactor. The reactor comprised a single stage
continuous flow, counter current glass ozone contactor with a liquid depth of 2.65 m and a diameter
of 0.15m. Mean operating conditions were: 6.4 L/min liquid flow rate; 0.64 L/min gas flow rate, 0.1
gas-to-liquid ratio and transfer efficiency of >90 % (Owens, 2000).They reported that under these
conditions, the CT required for 2-log inactivation of Cryptosporidium (C. parvum and C. muris) was 12
times greater than that required for the same degree of inactivation of Giardia muris cysts. Their
data suggested that this ratio may change depending on the log inactivation. Table 6 below (Von
Gunten et al., 2003) illustrates the difference in inactivation kinetics among a range of different
microorganisms including C. parvum.
Table 6. Inactivation Kinetics of Cryptosporidium and other pathogens or indicator organisms at pH 7 (adapted from Von Gunten et al, 2003).
Microorganism ko3 (l/mg/min) CTlag (mg.min/L) Temperature (°C)
E. coli 130 - 20
B. subtilis spores 2.9 2.9 20
Rotavirus 76 - 20
Giardia lamblia cysts 29 - 25
12a - 22
Giardia muris cysts 15.4 a - 25
C. parvum oocystsb 0.84 0.83 20 a= estimated value. Data derived from excystation for C. parvum. k refers to the rate constant for ozone disinfection.
It is worth noting that Owens et al. (2000) obtained a similar value of kO3 for C. parvum, while Finch
et al (1993) reported a different rate using an animal infection assay. Owens et al (2000)
demonstrated the need for relatively high CT values to inactivate Cryptosporidia compared with
other organisms. B. subtilis spores were also resistant (Table 7).
Page | 17
Table 7. Inactivation of a range of microorganisms in natural river water from Ohio, based on simple CT values (after Owens et al., 2000).
Microorganism n pH Temperature (°C)
CT (mg.min/L) Log inactivation range
Bacillus subtilis spores 19 7.9 22.7 0.7-18.4 0-2.17
C. parvum oocysts 6 8.2 24.5 2.6-7.2 0.57-2.67
C. muris oocysts 7 8.4 23.6 0.1-11 0.36-2.56
Giardia muris oocysts 5 7.6 25.2 0.3-1.0 1.52-2.70
Poliovirus 1 9 8.1 25.0 0.2-2.5 1.43-3.85
HPC, PCA 6 7.1 15.5 0.3-6.3 0.74-2.16
HPC, R2A 6 7.1 15.5 0.3-6.3 2.10-3.36
TC, P-A broth 6 7.1 15.5 0.3-6.3 2.63-3.95
TC, colilert 6 7.1 15.5 0.3-6.3 2.29-4.10 TC = total coliforms (colilert MPN method or P-A broth MPN method); HPC = heterotrophic plate count on PCA medium or R2A medium.
1.4 Factors affecting ozone disinfection
Ozone reacts with other constituents often present in raw or processed drinking waters. Ozone
demand is associated with the following (USEPA 1999):
Reactions with natural organic matter (NOM) in the water to form aldehydes, organic acids,
and aldo- and ketoacids (Singer, 1992).
Organic oxidation by-products (BDOC).
Synthetic organic compounds (SOCs), some of which can be oxidized and mineralized under
favourable conditions, particularly where hydroxyl radical oxidation is the dominant
pathway, (as in advanced oxidation processes).
Oxidation of bromide ion forming hypobromous acid, hypobromite ion, bromate ion,
brominated organics, and bromamines .
Bicarbonate or carbonate ions (alkalinity), forming carbonate radicals (Staehelin et al., 1984;
Glaze and Kang, 1988) – this is important in AOPs where the radical oxidation pathway is
predominant.
It is thought that molecular ozone is primarily responsible for inactivation of micro-organisms within
most operational ozonation pH ranges (between 6-9) (Hunt and Marinas, 1999). However the
production of hydroxyl radicals (OH) from the degradation of ozone results in the presence of
another highly oxidative species that also has disinfecting capability. Other contaminants present in
the water may scavenge both dissolved ozone and hydroxyl radicals, therefore the purity of the
water prior to ozone addition has a significant effect on its disinfecting capacity (US EPA, 1999).
These water quality parameters are now discussed in more depth:
Temperature: as temperature increases, it is generally thought that the disinfecting power of ozone
increases (Langlais et al., 1990). Whilst increased temperature reduces the solubility and stability of
ozone, temperature also increases the rates of diffusion of ozone through the cell wall of
microorganisms, as well as increased rates of reaction with the microbe. The effect of temperature
on Cryptosporidium has been clearly shown. A decrease in the required CT product for a 2-log
reduction in protozoan infectivity has been observed with an increase in temperature that followed
Page | 18
an Arrhenius type model (where K is the temperature in Kelvin, C is the chlorine concentration and T
is the contact time):
The model indicates that the CT required for 2-log reduction in Cryptosporidium increases by a factor
of 4.2 with every 10 °C decrease in water temperature.
Ozone inactivation of Cryptosporidia is temperature dependent, because temperature affects the
solubility, stability and reactivity of ozone (Martinez et al, 2011). For example, Driedger et al (2001)
observed that the rate of C. parvum inactivation decreased with decreasing temperature between 1
and 20 °C. Oppenheimer et al (2000) also demonstrated a reduction in CT values with increasing
temperature (Figure 2). CT values recommended by USEPA for inactivation of Cryptosporidium
oocysts by ozone are given in Table 8 below:
Table 8. CT values (mg.min/L) for inactivation of Cryptosporidium oocysts by ozone (after IEPA 2011).
Log Inactivation
Temperature (°C)
≤1 5 10 15 20
0.5 12 7.9 4.9 3.1 2.0
1.0 24 16 9.9 6.2 3.9
2.0 48 32 20 12 7.8
3.0 72 47 30 19 12
Source: USEPA, 2006.
By modelling ozone inactivation constants for natural waters across a range of temperatures,
Oppenheimer et al (2000) derived an empirical temperature characteristics term (Ɵ) for C. parvum.
This was expressed by the following equation:
k2/k1 = ƟT2-T1
where T1 is temperature 1; T2 is temperature 2; k1 = coefficient of specific lethality at T1; k2 =
coefficient of specific lethality at T2. The coefficient of specific lethality is derived from the change in
log inactivation of a particular organism with change in contact time with a specific disinfectant
under given conditions e.g. Verhille et al., 2003.
The authors reported that a multiplier of 4.5 for the CT value would be required for every 10 °C
temperature increase for C. parvum. Interestingly, this was substantially greater than the
temperature impact upon Giardia disinfection with ozone, thus demonstrating organism-specificity.
Page | 19
Figure 2. Illustrating the temperature dependence of CT values for 2-log inactivation of C. parvum in natural waters at 3 °C (open triangles), 7 °C (filled triangles), 10 °C (open squares), 20°C (filled squares), 22 °C (open circles) and 25 °C (filed circles). Data source: Oppenheimer et al., (2000).
pH: changes in water pH changes the balance of available O3 and OH, with an increase in pH from 6
to 9 reducing the amount of O3 available for disinfection by a factor of 40 (Elovitz et al., 1999). This is
because ozone decomposes to other products more quickly in alkaline environments. Over this pH
range, the available OH remains roughly constant, therefore it is thought that pH increases are likely
to influence disinfection more than oxidation reactions with inorganic and organic compounds.
However, empirical observations do not always confirm this. Hunt and Marinas (1999) saw only an
8% difference in second order inactivation rate for E. coli at pH 6 and 8. Driedger et al. (2001) saw no
effect on the inactivation of Bacillus subtilis spores between pH 6 and 8. It does, however, appear
that susceptibility to ozone at different pH is species specific. However, substantial effects of pH on
ozone inactivation of Bacillus subtilis spores was noted by Cho et al. (2003) by applying a hydroxyl
radical scavenger to all treatments, the effect of pH was negated and all CT values were similar. They
therefore attributed changes to the presence of hydroxyl radicals rather than pH. Langlais et al.
(1991) show evidence for decreased inactivation of poliovirus type 1 and rotavirus with increased pH
whilst inactivation of Giardia muris cysts increased from pH 7 to 9. A clear understanding of the
objective for disinfection must therefore be understood in order to determine how effective the
ozonation pH will be on disinfection efficiency.
There are conflicting findings in the literature relating to the effect of pH on ozone inactivation of
Cryptosporidium. For example, while there was no overall significant effect of pH on ozone
disinfection, Ran et al (2010a) demonstrated that in a batch reactor at pH 6 vs. pH 9, 3 mg/L ozone at
a contact time of 5 minutes led to a reduction in oocyst extinction from 99-93% respectively. In
earlier studies, Farooq et al. (1977) reported pH to have minimal effect where there is no ozone
residual i.e. continuous gas bubbling. In contrast, pH has been shown for Giardia muris to be
significant in batch systems where the O3 residual is decreasing (Labatiuk, 1992). In a batch study,
Kim et al (2007) found that ozone decay rates ranked from lowest to highest in the order pH
6.5<7.5<8.5, reporting a higher exposure ratio of hydroxyl radical to ozone as pH increased and
therefore indicating that pH effects on decay rates and production of hydroxyl radicals was a likely
0
10
20
30
40
50
60
70
80
90
100
0 10 20 30 40 50
Pe
rce
nti
le
CT value (mg min L-1)
Page | 20
cause of an zone-pH interaction. Oppenheimer et al. (2000) also observed a deviation from a linear
regression of pH vs. ozone inactivation of C. parvum at pH less than 6.5, indicating some inhibition of
disinfection.
Suspended solids & other ozone scavengers: Turbidity is important because it is indicative of the
presence of particulates. Firstly, microorganisms tend to associate with particulates and therefore
higher turbidities often correlate with higher pathogen loadings. Secondly, microorganisms can be
shielded from disinfectants when attached to particles or flocs, and particulates also exert a
disinfection demand leaving a smaller proportion of the transferred ozone to inactivate the
microbial target (Winward et al, 2008). This has been demonstrated for cell-associated poliovirus
and coxsackievirus at applied ozone doses of 4.1 and 4.7 mg/L respectively for 30 seconds. For un-
associated viruses, these were inactivated by the application of 0.08 mg/L of ozone for 10 seconds.
E. Coli inactivation has been shown to be reduced by the presence of an organic scavenger (tert-
butanol) (Hunt and Marinas, 1997). However, the nature of the material present has a big impact on
the consumption of ozone, therefore a parameter such as turbidity does not always give a good
measure of how the efficacy of disinfection will change. Non-reactive inorganic minerals will not
consume large quantities of ozone, but both particulate and dissolved organic substances are highly
reactive with ozone.
Several studies have reported an inverse correlation between turbidity and inactivation of
Cryptosporidia (e.g. Oppenheimer et al, 2000; Falabi et al 2002; Ran et al., 2010). Ran et al (2010)
evaluated the ozone-induced reduction in viability of Cryptosporidium (determined by fluorigenic
staining and microscopy) in distilled water laboratory batch reactors supplemented with different
concentrations of Kaolin to provide turbidity. The authors observed a decrease from 99.2 % to 86.2
% when turbidities increased from 0.1-20 NTU at a contact time of 7 minutes (Appendix 1).
Oppenheimer et al (2000), noted a significant difference in ozone inactivation in one set of four
replicate water samples in of natural waters and attributed this to a difference in turbidity of these
particular samples (1 vs. 11 NTU). Across the whole study, thought to represent water typologies
representative of around 65 % of those found in the US, high turbidity (>10) was correlated with less
effective ozone disinfection.
Dissolved organic matter has been shown to significantly affect the ozone inactivation of
Cryptosporidium in laboratory model systems in which humic acids were added to distilled water
batch systems, reducing the rate of inactivation from 84.9 % at 2 mg/L to 62.1 % at 10 mg/L (Ran et
al, 2010). Studies have also indicated a specific association between the number of oocysts present
and the ozone demand (Peeters et al., 1989); however Wohlsen et al., (2007) did not find this to be
the case. The presence of contaminants other than the target microorganisms may consume ozone
and reduce the disinfection capacity of the water (Dietrich et al., 2007). For example, viruses that are
associated with cells, or parts of cells, are protected from inactivation by ozone (Emerson et al.,
1982).
Strain related differences: C. parvum (found in over 150 mammalian species including humans) and
C. hominis (primary reservoir: Humans) are considered to be the most human and animal-health
relevant species of Cryptosporidium. There are eleven species in total, with differing primary hosts
and characteristics (Carey et al., 2004). Even subtle differences in oocyst size and shape may lead to
variations in the propensity to react with ozone, however there is little research on the subject and
Page | 21
compared with water quality factors, these are likely to play a limited role in mediated required
dosages of ozone for disinfection. Owens et al (2000) reported similar inactivation of C. parvum and
C. muris oocysts using ozone, although C. muris cysts were slightly more resistant than those of C.
parvum. Oppenheimer et al (2000) noted that at 20-25 °C, the Ct requirements of C. parvum were 7-
10 times higher than that described in the US Surface Water treatment Rule (USEPA, 1991) for
Giardia lamblia. This rose to 22 times greater at 10 °C. Therefore while intra-genus differences are
likely to be insignificant, reliance on disinfection kinetics for other organisms is not recommended.
In addition to differences related to genera, species and strains, Corona-Vasquez et al (2002)
reported that different “lots” of oocysts showed different levels of inactivation. Studies have
illustrated differences in the inactivation response of different ages of oocysts (Bukhari et al, 2000;
Driedger et al., 2001), which may have explained some of the findings.
It is therefore critical to know what is in the water to understand ozone doses and contact necessary
to achieve a specific water quality objective.
1.5 Ozone dosing and conditions
Under atmospheric conditions, ozone is an unstable gas that quickly decomposes to oxygen and has
a half-life of only 20 minutes at room temperature. Ozone is equally unstable when dissolved in
water and quickly breaks down to oxygen and a range of other products dependent on the water
quality (von Gunten, 2003b). In water, ozone’s half-life varies from seconds to hours and is a
function of the water pH, alkalinity and natural organic matter (NOM) concentration in the water.
Due to its instability, ozone must therefore be generated on site for immediate use. In addition,
ozone does not provide a residual disinfecting capacity so is not able to be used for this purpose and
an additional disinfectant much be added to the water if a residual is required (for example, using a
chemical such as chlorine), albeit usually at a lower dose.
When applying ozone at a WTW facility, there are four requisite components: (1) oxygen gas feed
system (either air or pure oxygen); (2) ozone production and delivery, (3) an ozone contactor and (4)
ozone off gas destruction.
(1) Gas feed: Ozone can be produced from air, oxygen or a combination of the two. A higher yield of
ozone is produced when pure oxygen is used as the feed for the ozone generator but it comes at
higher cost because the oxygen must be generated on site or purchased in a pure form. Yields of
ozone are relatively inefficient, typically 3.5% (by weight) from air and up to 14% from pure
oxygen gas (US EPA, 1999). For all oxygen sources, the gas source needs to be cleaned and dried
prior to ozone generation. This requires significant pre-treatment of the gas before it enters the
ozone generator, particularly when air is used. Typically, air is compressed, filtered, dried and
regulated prior to entering the ozone generator.
(2) Ozone generator and delivery: Ozone is formed when oxygen is broken down into atomic oxygen
radicals. In turn, these radicals may react with oxygen molecules to form ozone. The formation
of oxygen radicals requires high energy (493-683 kJ/mol) and the processes that are used to do
this are those that are able to form high energy electrons, usually from the application of high
voltage. This includes plasma corona discharge, chemonuclear sources and electrolytic
processes. Corona discharge is the most widely used process for large scale production of ozone.
Here, an electrical discharge is passed from a high potential electrode to a grounding electrode
across air or pure oxygen gas that has been dried and filtered (Langlais et al., 1990). Two
Page | 22
orientations are seen for corona discharge ozone generation, namely as concentric cylinders or
as parallel plates. Parallel plates only tend to be used in small scale ozone generation systems.
The concentric cylinders in most commercial systems look similar to conventional fluorescent
tube light bulbs and are composed of a high potential electrode encased in a glass dielectric
sleeve. The tube is then placed within a stainless steel grounding electrode with the gas passing
between the two electrodes (US EPA, 1999). Over 80% of the energy applied to ozone
generators is lost as heat, therefore this heat must be quickly dissipated to prevent quickened
degradation of ozone and overheating of equipment. Water is therefore circulated between the
electrode cells. The frequency of the power supplied to the generator impacts on the efficiency
of the ozone generation, with an increase in frequency resulting in an increase in efficiency.
However, higher frequencies tend to result in more heat generation and more complicated
power arrangements. In the past, low frequency systems were most common (50-60 Hz), put
medium (<1000 Hz) and high (>1000 Hz) systems have now become more popular as
advancements in electrical robustness have evolved.
(3) Ozone contactor: Ozone is contacted with water either via bubble diffusers, injector dissolution
or turbine mixers. Bubble diffuser systems are the most commonly applied contacting system
due to their relative simplicity, high ozone transfer rates (typically 85- 95%) and flexibility. Here,
water is passed through a baffled chamber in deep tanks (5.4-6.7 m) in either a co-, counter- or
combined co/counter current direction to the diffused ozone gas. Other ways to dissolve the gas
in the water are via injection of ozone into water under negative pressure. This may be directly
into the main flow (for small scale systems) or through a sidestream that is pressurised. The
sidestream is then mixed with the bulk flow under high turbulence. After the ozone is
distributed, a contactor must be provided to enable the appropriate CT to be delivered. Turbine
mixers are the final way by which ozone can be rapidly distributed into the flow. Turbine mixers
require high energy (4.8-5.9 kWh per kg of ozone transferred) but are able to achieve over 90%
ozone transfer. The main advantages and disadvantages of each type of ozone contacting
system is summarised in Table 9.
Table 9. Advantages and disadvantages of ozone mixings systems. Adapted from USEPA (1999).
Advantages Disadvantages
Bubble diffusion - No moving parts - Low hydraulic headloss - Simple to operate
- Deep contactors needed - Bubble channelling - Maintenance required of
diffusers, piping and gaskets
Injection - No moving parts - Effective ozone transfer - Shallow contactors needed
- High headloss from static mixers needed
- More difficult dosing flexibility
- Complex operation and high cost
Turbine mixing - High ozone transfer - Shallow contactor needed - Aspirating turbines able to
re-use off-gas from other chambers
- No filter clogging concerns
- High energy required - Maintenance of motor and
turbine
Page | 23
(4) Off gas destruction: the final component for any ozonation system is off-gas removal. As ozone is
a very toxic gas and because not all gas is transferred to the water, residual ozone must be
destroyed to prevent accumulation of ozone to harmful levels in working environments. Off gas
removal is carried out using heat, catalysts or a combination of the two.
The amount of applied ozone and the contact time required for disinfection is highly dependent on
the water quality being treated and the treatment target required. Typically applied ozone doses
range from 1-3.5 mg/L ozone, with most applications between 1.5-2.5 mg/L. Specific case study
examples for full-scale ozonation systems being used primarily for disinfection are shown in Table
10. Investigation of these applications of ozone show that tools such as online UV254, CT and CFD
modelling are all being used in order to optimise ozone doses for maximum pathogen control
(Bouland et al., 2004; Zhang et al., 2008; Audenaert et al., 2010).
Table 10. Ozonation case study sites for potable water disinfection.
Case study Ozonation procedure
Objective of ozonation
Typical ozonation conditions
Other information
A. H. Weeks Water Treatment Plant, Windsor, Canada Mazloum et al. (2004)
Pre ozone Cryptosporidium and Giardia control
1-3.5 mg/L - Improvements in coagulation observed
- Increase from 1 to 2 log reduction of Cryptosporidium and Giardia throughout year
Kluizen WTWs, Belgium Audenaert et al. (2010)
Post ozone Disinfection Taste & odour removal Biostabilisation
2.5 mg/L - UV absorbance (at 254 nm) used to help control ozone dosing on-line
DesBaillets WTP, Montreal, Canada Zhang et al. (2008)
Post ozone Disinfection 2.2 mg/L - CFD used to model disinfection efficacy – strong agreement between observed and modelled results
- Areas of short circuiting identified - CFD used for the first time as for an ozone residual monitoring strategy
Orly WTWs, Paris, France Bouland et al. (2004)
Pre and Post ozonation
Disinfection Taste & odour removal Biostabilisation
Minimum CT set at 1.6 mg/min/L 11-20 minutes contact time
- Authors suggest that the CT should include contact with downstream processes (e.g. GAC) due to higher levels of bromate formation observed on site.
1.6 Disinfection by-products
One of the key challenges faced when using ozone as an oxidising agent is the formation of
disinfection by-product (DBPs) compounds. DBPs are an unwanted consequence of adding strong
chemicals to water, and can be a range of hazardous compounds. Disinfection by-products (DBPs)
form as a result of reactions between chemical disinfectants, organic and inorganic matter (NOM
and IOM respectively) in water during the water disinfection process (Figure 3). Predominantly a
problem in drinking water treatment because of the direct risk of human consumption of water
containing harmful compounds, DBP formation could also be an issue in wastewater treated for
Page | 24
reclamation or where drinking water sources are significantly influenced by WWTP effluent (Wert et
al., 2007; Muller et al., 2012). Some DPBs associated with the halogens present either in raw water
or added for disinfection purposes, are known to be hazardous to health and can also be toxic to
aquatic systems. However, there is a requirement to protect human health from waterborne
disease, hence disinfection is essential. It is therefore important to achieve a balance between
effective disinfection and concentrations and types of DBPs formed.
There are two approaches to reduce the formation of DBPs during drinking water treatment. The
first is to reduce the presence of DBP precursors (principally NOM and bromide) prior to applying a
chlorine based disinfectant. This can be achieved through the use of combinations of chemical and
physical treatments (Chin et al, 2005), and may include ozonation. The second approach is to use a
non-halogen primary disinfectant (such as ozone) followed by the addition of a chlorine residual.
This reduces the amount of chlorine required and likewise the amount of chlorine-associated DBP
formation. However, as has already been noted, ozone is also known to produce a range of other
DPBs.
Figure 3. Schematic diagram of the reaction of organic and inorganic DBP precursors with disinfectants to form regulated and emerging DBPs (After Krasner, 2009).
Page | 25
1.6.1 Types and concentrations of DBP formed with ozonation
The formation of ozonation by products has been well documented and DBPs (Table 8) form when
disinfectants react with organic matter, bromide, iodide and anthropogenic pollutants (Richardson
and Postigo, 2012). DBPs may be organic (e.g., assimilable organic carbon (AOC), aldehydes,
carboxylic acids, and ketones) and inorganic (e.g., bromate). Nitrogen-containing DBPs, such as N-
nitrosodimethylamine (NDMA) which can form during the ozonation of the tolylfluanid fungicide
metabolite N, N-dimethylsulfamide (Schmidt and Brauch, 2008) are thought to be generally more
genotoxic and cytotoxic than those without nitrogen (Muller et al., 2012).
A comprehensive assessment by Richardson et al. (1999) identified the presence of hundreds of
halogenated (halo- alkanes/alkenes, aldehydes, ketones, dicarbonyls, acetonitriles, acids, alcohols,
nitromethanes) and unhalogenated (aldehydes, ketones, dicarbonyls, carboxylic acids, nitriles and
aldo and keto acids) DBPs from ozone, ozone-chlorine, ozone-chloramine samples, typically at very
low concentrations (ng-g/L ranges). When ozone is used alone, no halogenated DBPs will usually
form. However when ozone was dosed in addition to chlorine based disinfectants a number of
halogenated DBPs were formed at a concentration significantly higher than that formed with the
chlorine disinfectant alone. At present, the presence of these compounds do not present a
regulatory risk, but as future water quality regulation becomes stricter in the coming years, it is
important to understand the presence, concentration and toxicity of these DBPs.
While many organic and inorganic ozonation disinfection/ oxidation by-products have been
identified (Table 11), bromate is generally considered to be of greatest concern (von Gunten, 2003)
and aldehydes are also important although they are not currently regulated (Silva et al., 2010; Table
12). Where bromide is present in raw waters, bromo-organic by-products can form during ozonation
(Figure 2). Ozone and bromide react to form hypobromous acid, which then reacts with organic
matter to produce the DPBs. These include: bromoform, bromopicrin, dibromoacetonitrile,
bromoacetone, bromoacetic acid, bromoalkanes, bromohydrins, but most tend to be present in low
concentrations (von Gunten, 2003).
Iodate is formed by the ozone-oxidation of naturally occurring iodide and is of no toxicological
concern (von Gunten, 2003). Bromate is the only specific DPB of ozone regulated in drinking water in
a range of countries, including the UK and the US, with a limit of 10 μg/L (Table 12).
Aldehydes, primarily formaldehyde, acetaldehyde, glyoxal, and methylglyoxal are formed as a result
of the oxidation of organic matter from wastewater (Melin and Odegaard, 2000; Huang et al., 2005;
Silva et al., 2007). Silva et al (2010) determined concentrations of ozonation by products generated
from anaerobic (USAB) wastewater effluent in a bench scale continuous ozonation system.
Aldehydes were present at up to 187 µg/L and glyoxal up to 46 µg/L. These are regulated in only a
few countries, and then usually for specific compounds such as formaldehyde and
trichloroacetaldehyde (Table 12). Contact time did not appear to affect the degree of aldehyde
formation, for contact times between 5 and 15 minutes and doses of 5, 8 and 10 mg O3/L. In
contrast, Nawrocki et al (2003) found strong contact time dependency of formaldehyde and
acetaldehyde during the ozonation of drinking water.
Weinberg et al (2002) carried out a National Occurrence study, evaluating DBP concentrations across
12 drinking water treatment plants in the US in which sites using all four major disinfectants
Page | 26
(chlorine, chloramines, ozone, and chlorine dioxide) were covered. It is notable that the highest
concentration of dichloroacetaldehyde occurred at a plant using chloramine and ozone disinfection.
Therefore, although the use of alternative disinfectants minimized the formation of the four
regulated THM, some unregulated DBPs were present in higher levels than where traditional
chlorine disinfection was applied.
Table 11. Some key DBPs generated by Ozone and their Health Effects (adapted from Lipmann, 2009).
Compounds Group
Oxyhalides Aldehydes
Carboxylic Acids
Health Effects Animal, possible human carcinogen (Bromate) Limited knowledge on toxicity of chlorate
Can be carcinogenic at high levels but low concentrations present in water contribute little to calculated cancer risk. Questionable importance as carcinogens via ingestion (formaldehyde and acetaldehyde). (Lippmann, 2009)
Possible developmental toxicity (Moudgal et al, 2000)
Specific DBPs Bromate Formaldehyde 2-Methylpropanoic acid
Chlorate Acetaldehyde Butanoic acid
Chlorite Cyanoformaldehyde Pentanoic acid
Glyoxal Hexanoic acid
Methylglyoxal Heptanoic acid
Propanal Octanoic acid
Butanal Oxalic acid
Pentanal Malonic acid
5-Keto-1-hexanal Succinic Acid
Trans-2-Hexanal
Page | 27
Table 12. DBP Regulations worldwide (adapted from Hrudley and Charrois, 2012).
DBP Guideline value (µg/L) by Country/Agency
WHO EU USEPA Australia Canada Japan
Chloroform 300 - - - - 60
Bromodichloromethane 60 - - - - 30
Chlorodibromomethane 100 - - - - 100
Bromoform 100 - - - - 90
Monochloroacetic acid 20 - - 150 - 20
Dichloroacetic acid 50 - - 100 - 40
Trichloroacetic acid 200 - - 100 - 200
Trichloroacetaldehyde - - - 20 - -
Formaldehyde - - - - - 80
Bromate 10 10 10 20 10 10
Chlorate 700 - - - 1000 -
Chlorite 700 - 1000 800 1000 -
Dichloroacetonitrile 20 - - - - -
Dibromoacetonitrile 70 - - - - -
N-Nitrosodimethylamine 0.1 - - 0.1 0.04 -
Total THMs - 100 80 250 100 100
HAA5 - - 60 - 80 -
Cyanogen Chloride - - - 80 - -
However, undoubtedly the DBP of most concern at present when using ozone is bromate. Due to its
power as an oxidising agent, ozone is able to oxidise bromide to bromate. Bromate is a known
carcinogen in mammals and is therefore strictly controlled at 10 g/L in Europe in drinking water
supplies. There are a number of factors that influence bromate formation. These are:
Bromine concentration: given that bromide is oxidised by ozone to bromate, an increase in
bromide leads to an increase in bromate for a constant ozone dose and contact time
(Legube et al., 2004). Conversion of bromide to bromate is relatively high (typically between
10-50 %), with higher conversions observed in summer compared with winter (Song et al.,
1996). Typical concentrations of bromide in natural waters usually range from 30-200 g/L,
with an average of 100 g/L (Amy et al., 1993), however this can be greater than 500g/L
(Legube et al., 2004). Amy et al. (1993) approximated that up to 30 g/L of bromate can
form from a bromide concentration of 100 g/L, which is a similar conversion ratio to that
seen by others. High bromide in water sources can result from road run-off, ingress by saline
water and from the dissolution from sedimentary rocks (Magazinovic et al., 2004; Butler et
al., 2005). General rules for bromide containing waters are that those containing <20 g/L of
bromide do not present a problem for bromine-derived DBP’s, whilst waters containing >100
g/L of bromide are likely to cause significant bromate problems (von Gunten 2004a). Data
from full-scale water treatment plants show that other chemical factors than just bromine
concentration are important when understanding bromate formation rates (Table 13). These
factors are listed below.
Page | 28
Table 13. Bromate formation at full scale WTWs in Europe and N America. Adapted from von Gunten (2003b).
Location Number of plants Bromide range (g/L) Bromate range (g/L)
France 42 12-658 <2-19.6
Germany 4 30-150 <1-12
Switzerland 86 5-50 <0.5-20
USA 24 2-180 0.1-40
Temperature: Higher temperatures increase the rate of bromate formation as a result of
increased reaction kinetics and because there is a commensurate increase in the acidity
constant which promotes the formation of an important precursor of bromate (Legube et
al., 2004). The effect of temperature is more pronounced at higher ozone doses. For
example, Galey et al. (2004) observed that at an ozone dose of 1 mg/L the bromate
formation was 8 g/L at both 5 and 24 °C whilst at 2.5 mg/L the bromate formation was 22
g/L at 5 °C and 37g/L at 24 °C.
pH: As the pH of the water is increased during ozonation, more bromate is formed. At low
pH, more hypobromous acid remains fully associated with protons, which prevents the
formation of an important intermediary (BrO-) in the bromate formation pathway. At higher
pH, the formation of this compound is favoured. Hydroxyl radical formation is also promoted
at high pH due to the increased concentration of hydroxyl ions present (Song et al., 1997;
Siddiqui et al., 1998). Bromate formation has been shown to increase from 10 g/L at pH 6.5
to 50 g/L at pH 8.2 (Legube et al., 2004) whilst Krasner et al. (1994) observed a 60 %
decrease in bromate formation for each drop in pH unit. The ozonation pH is the best way of
controlling bromate formation at a WTW (Ozekin and Amy, 1997). This must be countered
by the increased formation of brominated organic compounds as the pH is reduced (USEPA,
1999) and because high alkalinity waters may require too much acid to reduce the pH to be
cost effective (von Gunten, 2003a).
Alkalinity: The presence of inorganic carbon (IC) species increases bromate formation
because both carbonate (CO2-) and bicarbonate (HCO3
-) species can form the carbonate
radical (CO3-•) as a result of oxidation by hydroxyl radicals (von Gunten, 2003a). Once the
carbonate radical has been formed, this can convert hypobromite into the hypobromite
radical (BrO-•) and then bromate (Kim et al., 2004).
Ammonia concentration: The presence of ammonia in water acts as a scavenger of
hypobromous acid (HOBr) during ozonation, an important intermediate in the formation
pathway of bromate (von Gunten, 2003a). HOBr reacts with ammonia to form bromamine
compounds, which, in turn, can be converted back to bromide through oxidation by ozone.
Ammonia can therefore remove a significant intermediary from the bromate formation path
and reduce the amount of bromate formed (Song et al., 1997). Ammonia may be present
naturally in waters to be ozonated, or alternatively can be added prior to ozonation as a
bromate prevention strategy. The addition of 1.5 mg/L ammonia addition can reduce
bromate formation by around 5 g/L when applied to water containing 100 g/L Br- under
constant conditions (Ozekin and Amy, 1997). This reduction, although small, may be critical
for those WTW where bromate levels are around the maximum permitted concentration.
Page | 29
However, this must be tempered by the fact that above a certain ammonia concentration,
the addition of more ammonia has no further effect on bromate reduction. Therefore, for
waters that contain naturally high to medium concentrations of ammonia, the addition of
further ammonia may offer no further benefit (von Gunten, 2003b). Furthermore, any un-
removed ammonia may act as a nutrient for nitrifying bacteria once in distribution (USEPA,
1999a).
Transferred ozone dose and contact time: As previously discussed, the efficiency of any
disinfectant may be characterised by the CT factor (USEPA, 1999b). The combined impact of
the concentration of applied ozone and residence time of the ozone in the reactor is an
important parameter in the formation of bromate. The relationship between bromate
formation and CT follows a linear function, with an increase in CT leading to an increase in
bromate formation (von Gunten and Hoigne, 1996 and Legube et al., 2004). Due to its low
solubility, typical residual concentrations of ozone found at WTW are in the range 0.1-1
mg/L. In order to achieve 99 % inactivation of Cryptosporidium oocysts typical contact times
in the range of 4-18 minutes are applied giving CT of 2.4-10 mg min/L (USEPA, 1999a).
However, CT is dependent on temperature and the log inactivation of microorganisms
required. For example at 13 °C, a 3-log inactivation of Cryptosporidium requires a CT of 22
mg min/L whilst at 22 °C a similar inactivation requires a Ct of 8 mg min/L (Galey et al.,
2004). Any change in CT to control bromate formation must therefore also ensure that
adequate disinfection is maintained.
Empirical models have been developed that predict bromate formation based on these water quality
and ozonation parameters. The variables important for bromate formation are those mentioned
previously (i.e. bromide, DOC or UV254, pH, O3 dose, NH3, alkalinity and temperature). The
relationship of each of these variables to the output bromate concentration is then found
experimentally by fixing all variables but one. The change in bromate formation is then observed
with the random variable. By carrying out multiple linear regression (MLR) analysis on the data (or
log transformed data), the cumulative relationship and significance of each of the variables can be
found. MLR was first applied to bromate formation by Ozekin (1994) and, to date, most bromate
formation models using MLR have been of the form:
log Y = b0 + b1 log x1 + b2 log x2 + b3 log x3…….+ bn log xn
where Y is the dependent variable, xi is an independent variable and bi is the regression coefficient.
The following example shows the regression model for bromate formation from Song et al. (1996):
log [BrO3-] = -6.11 + 0.880 log[Br-] – 1.180 log[DOC] + 5.110 log[pH] + 1.420 log[O3] + 0.270 log[t] –
0.180 log[NH3-N] + 0.180 log[IC]
A range of bromate formation models found in the literature of the form shown above are included
in Table 14. An alternative form of the equation has been developed by Ozekin and Amy (1997):
log[BrO3-] = -3.361 + 1.136 log[Br-] – 1.267 log[DOC] + 0.249 [pH] + 1.575 log[O3] + 0.006 [t]
DOC has been replaced with UV254 in some of the models (Sohn et al., 2004). This is advantageous
because UV254 is an easier and more robust measurement to make on-line and UV254 is more suitable
for application to waters where staged ozonation is deployed or for post ozonation because UV254 is
Page | 30
significantly reduced during the first ozonation process, however the DOC of the sample stays
approximately the same. During intermediate and staged ozonation, less ozone is consumed by the
DOC compared to un-ozonated water of a similar DOC. The effect of this being an increase in the
bromate formation potential. DOC based regression models should therefore only be applied to one-
off ozonation situations. A number of the regression models also include ammonia. This gives
flexibility to water utilities depending on whether they add ammonia as a bromate control measure
or routinely measure ammonia as a water quality parameter (Ozekin and Amy, 1997).
The models show that the parameters that most affect bromate formation are of the following order
for increasing bromate formation: pH > O3 dose > Br- > IC > time and for decreasing bromate
formation: DOC > NH3-N. When the models have been applied to similar raw waters to those with
which the models have been developed (internal validation), good correlation has been seen
between the observed and predicted bromate concentrations. For example, Song et al. (1996) had
an average R2 value of 0.93 for validation of bromate models on water sources that were used to
develop the model for predicted against measured bromate concentrations. Similarly good
correlation was seen by Siddiqui et al. (1994) and Ozekin and Amy (1997) with R2 values of 0.98 and
0.91 respectively. However, increased error becomes apparent when the models have been
validated with external data. Nevertheless, these models show the key way in which water quality
parameters influence bromate formation and if the time can be taken to develop specific formation
models for specific waters, a high degree of accuracy on bromate formation can be ascertained.
While there is continuing debate about the issue, there is evidence to show that the risk of infection
by waterborne disease can outweigh that from DBPs. For example, the risk of infection with C.
parvum was considered to outweigh that of getting renal cancer due to consumption of bromate-
containing drinking water resulting from ozonation of waters containing bromide (Havelaar et al.
2000). Application of ozone to suitable waters or effluents has the potential to improve human
health by maintaining protection against pathogens while removing or reducing the requirement for
chlorine application (Wert et al 2007) and consequently reducing the formation of key
trihalomethane and haloacetic acids present in drinking water or entering the environment.
Page | 31
Table 14. Regression coefficient value for bromate formation from various locations. Adapted from Jarvis et al. (2007).
Regression coefficient values for listed variables from bromate prediction models (number of variables included is dependent on particular model)
Boundary conditions Notes Source
Constant
Br
g/L)
DOC (mg/L)
pH O3*1
(mg/L) t*2
(mins) NH3-N
(mg/L N) UV254 (cm-1)
Alkalinity (mg/L as CaCO3)
Temp (°C)
-5.810 0.730 -1.260 5.820 1.570 0.280 - - - - 70 ≤ Br- ≤ 440 1.1 ≤ DOC ≤ 8.4 6.5 ≤ pH ≤ 8.5 1.1 ≤ O3 ≤ 10.0
1 ≤ t ≤ 120
Model developed from 10 raw waters. For raw waters with no ammonia. Limited to 20 °C.
Ozekin (1994)
-5.788 0.730 -1.300 5.790 1.590 0.270 -0.033 - - - 70 ≤ Br- ≤ 440 1.1 ≤ DOC ≤ 8.4 6.5 ≤ pH ≤ 8.5 1.1 ≤ O3 ≤ 10.0
1 ≤ t ≤ 120 0.02 ≤ NH3-N ≤ 3.0
Model developed from 10 raw waters. For raw waters with ammonia. Limited to 20 °C
Ozekin (1994)
-6.924 0.960 - 5.680 1.307 0.336 - 0.623 -0.201 *3 70 ≤ Br- ≤ 440 6.5 ≤ pH ≤ 8.5 1.1 ≤ O3 ≤ 10.0
1 ≤ t ≤ 120 0.010 ≤ UV254 ≤ 0.280
13 ≤ Alkalinity
Developed from Ozekin (1994). For raw waters with no ammonia. Limited to 20 °C.
Sohn et al. (2004)
-7.080 0.944 - 5.810 1.279 0.337 -0.051 0.593 -0.167 *3 70 ≤ Br- ≤ 440 6.5 ≤ pH ≤ 8.5 1.1 ≤ O3 ≤ 10.0
1 ≤ t ≤ 120 0.02 ≤ NH3-N ≤ 3.0
0.010 ≤ UV254 ≤ 0.280 13 ≤ Alkalinity ≤ 316
Developed from Ozekin (1994). For raw waters with ammonia. Limited to 20 °C.
Sohn et al. (2004)
-6.110 0.880 -1.180 5.110 1.420 0.270 -0.180 - 0.180 [says IC] - 2 ≤ BrO3-
100 ≤ Br- ≤ 1000 1.5 ≤ DOC ≤ 6.0 6.5 ≤ pH ≤ 8.5 1.5 ≤ O3 ≤ 6.0
0 ≤ t ≤ 30 0.02 ≤ NH3-N ≤ 3.0 1 ≤ Alkalinity ≤ 216
Model developed from 4 different model waters. Limited to 20 °C
Song et al. (1996)
Page | 32
Regression coefficient values for listed variables from bromate prediction models (number of variables included is dependent on particular model)
Boundary conditions Notes Source
Constant
Br
g/L)
DOC (mg/L)
pH O3*1
(mg/L) t*2
(mins) NH3-N
(mg/L N) UV254 (cm-1)
Alkalinity (mg/L as CaCO3)
Temp (°C)
-4.267 0.040 -1.080 4.700 1.120 0.304 - - - 0.580 70 ≤ Br- ≤ 440 1.1 ≤ DOC ≤ 8.4 6.5 ≤ pH ≤ 8.5 1.1 ≤ O3 ≤ 10.0
1 ≤ t ≤ 120
Galey et al. (1997)
-2.824 0.610 -0.740 2.260 0.640 - - - - 2.030 250 ≤ Br- ≤ 1500 3.0 ≤ DOC ≤ 7.0 6.5 ≤ pH ≤ 8.5 1.5 ≤ O3 ≤ 17.5 20 ≤ Temp ≤ 30
Model developed from 5 surface and ground waters
Siddiqui et al. (1994)
*1Utilised/transferred ozone; *2Time in ozone contactor; *3Temperature correction factor can be applied for variations in temperature: [BrO3]@tempT = [BrO3]@temp20 °C (1.035)T-20
Page | 33
2.0 Oxidation of contaminants in water Compounds present in water can react with ozone directly or indirectly through OH radicals (von
Gunten, 2003b). These two different reaction pathways lead to different oxidation products and are
controlled by different types of kinetics. Many of the rate constants for the direct reaction with
ozone appear to be low and typically in the range 1.0 – 103 per M/s (Gottschalk et al., 2000). One of
the key molecule characteristic which controls reactivity of a compound with ozone is its charge or
polarity. For example, ionised or dissociated forms of organic compounds will react with ozone much
faster than a neutral (non-dissociated) form.
The first step in the indirect reaction pathway is the decay of ozone which can be accelerated by the
presence of OH- ions (Gardoni et al., 2012). The radical pathway is very complex and can be
influenced by many substances. The mechanism of the indirect pathway consists of initiation, chain
reaction and termination steps. Substances which convert OH into superoxide radicals promote the
chain reaction but if OH react with a compound, and this does not lead to the formation of a
superoxide radical, the chain reaction is terminated and the ozone decay inhibited. Such compounds
are often called OH scavengers and these can include carbonate and bicarbonate ions, humic acid
(which can also act as a promoter) and phosphate (Stahelein and Hoigne, 1985).
Direct ozonation is usually the most important oxidative reaction if the radical reactions are
inhibited due to the lack of initiating compounds to begin the chain reaction or due to the presence
of too many radical scavengers (Gottschalk et al., 2000). The direct pathway normally dominates
under acidic conditions (pH <4) and changes to the indirect pathways above pH 10. Both pathways
will therefore play a role in most ground and surface waters (pH ~ 7). It is therefore clear that pH has
a great effect on the dominant reaction pathway, however it will also depend on the contaminants
present in the water or wastewater.
2.1 Destruction of organic and inorganic compounds using ozone
A number of compounds can be directly degraded by ozone. These include taste and odour
compounds (geosmin and methylisoborneol (MIB)), phenolic compounds and pesticides such as
atrazine (Crittenden et al., 2005). The importance of ozonation in the treatment of industrial
wastewaters targeting the degradation of dyes, pharmaceuticals and personal care products has also
grown in recent years. Ozone is also used for oxidation of organic macropollutants and its
application is used for bleaching of colour, increasing the biodegradability of organic compounds,
removal of THM precursors and reducing total organic halide formation potential or chlorine
demand (Langlais et al., 1991).
Several options exist to promote OH formation during ozonation. These include the addition of H2O2,
TiO2 and combining ozone with UV (von Gunten, 2003a). The O3/H2O2 process is perhaps the easiest
and cheapest option to enhance the process, but low reaction rates have been observed at pH 5 or
less (Matilainen and Sillanpää, 2010). Theoretically 0.5 moles of H2O2 is needed per 1 mole ozone
(0.354 kg H2O2/kg O3) but more ozone is usually required as it is more reactive with the background
organic matter. However care must be taken because excess ozone can have a scavenging effect on
the hydroxyl radicals formed. In addition, careful control of the H2O2 residual is needed because of
its higher stability when compared with ozone (Crittenden et al., 2005).
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Several studies have looked into the removal of NOM by ozone based AOPs and concluded that
these processes increase the removal of NOM but do not always decrease the production of THMs
when compared to normal ozonation (Matilainen and Sillanpää, 2010). Catalytic ozonation is a
promising technology for the effective removal of pollutants that are resistant to conventional water
treatment but the mechanisms of these processes are still unclear (Legube and Leitner, 1999; Ikhlaq
et al., 2013).
In all applications ozone is consumed not only in the oxidation reaction with the target compound
but also with the background water matrix (non-target demand) as well as through self-
decomposition. The overall ozone demand is water quality specific and may vary widely with season
and temperature and therefore existing applications on the same water or treatability studies are
the only ways of determining the dose requirements.
Optimum contact time can also vary widely depending on the intended ozonation objective and
again this is usually determined through treatability studies. Oxidation of easily oxidisable
compounds usually happens in a much shorter timeframe than disinfection. The contact time for
oxidation applications is usually far less important than the dose itself and this is mainly true for fast
ozonation reactions. However the contacting system can have an effect on the reaction rate.
Further, contact time does not necessarily have the same meaning for different types of reactors as
mixing intensities and rate of transfer can have a dramatic effect on the relative contact time
associated with a process.
2.1.1 Iron and manganese removal
One of the most important ozone applications in water treatment is the oxidation of iron and
manganese. It is not considered to be a very cost effective strategy in the US but it is successfully
applied in Europe in pre or inter- ozonation combined with conventional treatment processes
(Crittenden et al., 2005). Stoichiometric requirements are 0.43 mg O3/mg Fe2+ and 0.86 mg O3/mg
Mn2+ but often higher doses are required due to competing oxidation reactions that happen during
the treatment stage. Ozone can readily oxidise iron and manganese in groundwater and water of
low organic content. Uncomplexed iron is oxidised much faster than manganese as kinetics of iron
oxidation by ozone are very fast and therefore as a result, if high concentration of iron is present in
groundwater, low ozone doses lead to little manganese oxidation. The ozone doses required in such
environments are close to stoichiometric doses if no other scavengers (nitrites or sulphides) are
present. Excessive ozone doses (2.2 mg O3/mg Mn2+) will lead to the formation of permanganate
leading to pinkish colour of the treated water. The oxidation of manganese by ozone is less
dependent on pH than for other oxidants and ozone is more likely to offer kinetic advantages at low
rather than high pHs. An additional advantage of groundwater ozonation is re-oxygenation of the
water.
The degree of complexed iron oxidation by ozone is likely to be dependent on the pH and the nature
and concentration of the organic matter. Ozonation can also lead to the formation of stable iron-
organic complexes that cannot be subsequently further oxidised. Therefore it is important to remove
organic matter prior to ozone application. It was reported previously that humic substances can
successfully compete for ozone leading to higher doses required (Rakness, 2005). For example, for
synthetic water (alkalinity 150 mg/L; pH 8.5; [Mn] = 250 g/L) 0.5 mg O3/mg Mn as ozonation dose
has resulted in a filtered-water manganese residual of less than 30 g/L and higher ozone doses led
Page | 35
to the formation of permanganate. In another case for a synthetic water ([Mn] = 1 mg/L; pH 6.3; TOC
= 2-5 mg/L; 50-200 mg CaCO3/L), 75% manganese removal was achieved after application of ozone
dose of 0.2-0.7 mg O3/mg C in excess of that required for manganese oxidation alone. The excess of
ozone required was inversely proportional to the alkalinity and high levels of bicarbonate lead to
lower ozone requirements. For this reason, intermediate recarbonation has been used to improve
the efficiency of manganese oxidation by ozone (El Araby et al., 2009). Once the manganese is
oxidised to Mn(III) or Mn (IV) it can be removed by flocculation and sedimentation, rapid sand
filtration or multimedia filtration. Examples of iron and manganese removal by ozone are shown in
Table 15 below.
Table 15. Removal of iron and manganese in various case studies (adapted from Langlais et al., 1991).
Water source Average dose Contact time Process Efficiency
Groundwater
(average Fe = 50 g/L and Mn = 75 g/L)
0.1 mg/L 2 min + 2 min Fe: 80% removal Mn: >50% removal
River water
(Fe = 23-263 g/L and Mn = 20-300 g/L)
0.5 mg/L (max 1.9 mg/L)
3 min + 5 min Concentrations below detection limit
Reservoir water (pilot)
(average Fe = 600 g/L and Mn = 150 g/L;
TOC = 7.5 mg/L
0.8-1.3 mg/L not available Fe: 99% removal Mn: 85% removal
Reservoir water
(Mn= 30-550 g/L; TOC = 3.8-4 mg/L)
0.8-1.3 mg/L 2 min <50 g/L
If post-ozonation (for disinfection) is included in the treatment train without subsequent filtration,
manganese removal by pre- or inter-ozonation must be well controlled to prevent further oxidation
of residual manganese in the post-ozonation stage which could lead to the production of an
unacceptable pinkish colour in the water. Other inorganic compounds that can be removed by
ozone, usually during pre-ozonation, includes nitrite (NO2-) and H2S/S-2 (von Gunten, 2003a).
2.1.2 Oxidation of organic macropollutants
The removal of NOM can have a number of objectives, including removal of colour and UV
absorbance, an increase in biodegradable organic carbon ahead of biological treatment, reduction of
the DBP formation potential and direct reduction of DOC/TOC levels by mineralisation. The principle
oxidation by-products from NOM ozonation are aldehydes, ketones and carboxylic acids (Westerhoff
et al., 1999). Mineralisation of NOM is usually not achievable at full scale as this requires high ozone
as shown in Table 16.
Table 16. Ozone demand for oxidation of organic compounds.
Target objective Reported ozone demand Reference
Colour removal Up to 3 mg O3/mg C Langlais et al., 1991; Gottschalk et al., 2000
Increase in biodegradable carbon
1-2 mg O3/mg DOC Gottschalk et al., 2000
Reduction of disinfection by- 0.5-2 mg O3/mg DOC (to achieve Gottschalk et al., 2000
Page | 36
product formation potential 10-60% reductions)
TOC/DOC mineralisation 3 mg O3 mg/DOC (for 20% removal efficiency)
Gottschalk et al., 2000
When used in drinking water treatment, ozonation is usually positioned between settling/flotation
and rapid filtration or between rapid filtration and activated carbon filters or other post-treatment
units (Gottschalk et al., 2000). Because ozone interactions with NOM causes substantial changes in
the organic compound’s molecular structure, ozonation can be also applied prior to membrane
filtration to reduce membrane fouling (Van Geluwe et al., 2011).
Ozone doses of 1-3 mg O3/mg C lead to almost complete colour removal (Langlais et al., 1991) but
because humic substances can account for over 10 mg/L of DOC in many natural waters, the
required ozone doses are likely to be inhibitive due to cost and the likely high rate of DBP formation.
Other sources suggest that over 90% colour removal can be achieved with an ozone dose in the
range below 1 mg O3/mg DOC. If sufficient ozone dose is applied it can lead to a reduction in chlorine
demand. It has also been shown that beyond a certain threshold, the colour was difficult to remove
and further treatment is required (GAC filtration or sand filtration) or multistage ozonation is
necessary.
Topley (1987) reported the application of ozone in combination with a biological activated filter
process for the treatment of highly coloured water (up to 60 total colour units (TCU)). The ozone
dose in the full scale plant was designed to be 4.2 mg/L at full flow with a 5 minute contact time in
each of the four contact chambers and it was reported that substantial colour removal occurred at
an ozone dosage of approximately 2-3 mg/L with some additional colour removal occurring through
the filter (Topley, 1987).
2.1.3 Removal of micropollutants
Ozone is capable of destroying a range of volatile micropollutant compounds, in particular alkenes
and aromatic organics, using the conditions of ozone treatment applied in drinking water (Langlais et
al., 1991). In the past, micropollutant removal has not been usually a primary task for ozone but was
considered to be a positive side effect. However, due to ever lowering detection limits and stricter
regulatory requirements for more chemicals in drinking water, the interest in micropollutants has
grown in recent years. The oxidation of micropollutants by ozone is only an efficient process for
compounds that contain an amino group, an activated aromatic system or a double bond (von
Gunten, 2003a). Most micropollutants are poorly accessible to direct ozone attack and an advanced
oxidation process is usually required. Micropollutants are usually transformed to metabolites that
are often more polar in nature and smaller in molecular weight, but are rarely completely
mineralised, therefore it is essential to have a subsequent treatment unit (Gottschalk et al., 2000).
The following table (Table 17) summarises the expected removals of a range of micropollutants
when using ozone.
Table 17. Degree of removal of trace organics during ozonation in full-scale drinking water treatment plants (Gottschalk et al., 2000).
Substances Degree of removal (%) Comments
Taste and odour 20-90 source specific
MIB, geosmin 40-95 Improvement by AOP (O3/H2O2; O3/UV)
Alkanes <10
Page | 37
Alkenes, chlorinated alkenes 10-100 chlorine content important; AOP support oxidation
Aromatics and chloroaromatics 30-100 highly halogenated phenols are more difficult to oxidise
Aldehydes, alcohols, carbonic acids low typical products of ozonation, easily biodegradable
N-containing aliphatics and aromatics 0-50 AOP may increase oxidation rate
Pesticides 0-80 very substance specific; triazines require AOP
Polyaromatic hydrocarbons High, up to 100
2.1.4 Taste and odour control
The efficiency of ozone in the removal of taste and odour depends on the compounds causing the
issues and if these compounds are saturated structures, ozone can have little impact. The most
common taste and odour associated compounds are MIB and geosmin and these have a very low
reactivity with ozone. However, despite this, studies with natural waters have shown good removal
efficiencies of these compounds when using ozone. It is likely that ozonation is most effective in
waters that support the OH radical pathway. However, the action of ozone in natural waters is
variable and depends on the quality of organics present as well as the treatment conditions. It was
reported that the combination of ozone with H2O2 increased geosmin and MIB elimination by 35%
compared to ozone alone (Langlais et al., 1991). The best taste and odour control strategy includes
ozonation followed by filtration and adsorption. If combining ozonation with GAC adsorption, it is
important to control the ozone dose in order to prevent the formation of certain by-products such
as short chain aldehydes which are not very well adsorbed on GAC (Langlais et al., 1991). In most
cases only a partial oxidation of the compounds is necessary to eliminate taste and odour problems
(von Gunten, 2003a).
2.1.5 Particle removal
It has been observed for more than 30 years that pre-ozonation ahead of solid-liquid separation
processes can improve the removal of particles (Gottschalk et al., 2000). Pre-ozonation can have a
number of positive effects on coagulation and settling processes and these include a shift in particle
size distribution towards larger sizes, the formation of colloidal matter from dissolved DOC, a
decrease of coagulant dose and an increase in floc settling velocities. It has been suggested that a
certain critical concentration of organic material must be present in order to observe the enhanced
coagulation effects of ozone. Further the calcium concentration in raw water also impacts on the
coagulation effects of ozone. Ozone gas can be added either before or together with the coagulant
at dosages between 0.5-2 mg/L. The mechanisms involved appear to be rather complex and poorly
understood. The observed increase in solids removal are often quite variable from site to site (20-
90%) which suggests some very specific reactions are taking place from water source to water
source (Gottschalk et al., 2000). Positive effects of pre-ozonation are also seen seen for algae
removal. Ozone readily kills or lyses many types of algae and it has also been observed to enhance
the removal of algae by coagulation and settling. Ozonation of algae can lead to the release of
surface active polymers, which may aid aggregation but could also be a source for organic
disinfection by-products. Ozone can also be applied to inactivate zooplankton and actinomycetes
(Lin et al., 2012). A number of laboratory studies have reported the effect of ozonation on the
removal of cyanotoxins and it was shown that complete removal of the toxins can be achieved when
Page | 38
ozone is included in the treatment process (von Gunten, 2003a; Sharma et al., 2012). Ozoflotation is
a new process combining the physical phenomenon of flotation with the oxidising properties of
ozone and is usually used as a pre-treatment stage in order to reduce the treatment load.
3.0 Point of use disinfection of water supplies
3.1 Applications and dosing conditions
There are a range of commercially available point of use (POU) and point of entry (POE) ozonation
devices. In POU systems, ozone is added directly into the water that is being used, for example at the
tap. In POE systems, ozone is added to the water as it enters a property before it is distributed
throughout the premises. These devices have found application for drinking water treatment,
wastewater treatment, aquaculture and agriculture, electronics manufacture, food production and
pool water treatment. There is limited scientific literature available on the performance and
reliability of these devices and most of the below information has been taken from commercial
sources and, as such, limited validation of performance can be gleaned from this data.
POU and POE systems generate ozone in the same way as for full scale systems as applied at WTWs,
usually by passing high voltage electricity through air. Details of POU systems found in the literature
are shown below:
Ozone Pure Water Inc: http://www.ozonepurewater.com/ozone.htm#2
Ozone is generated by high voltage electricity passed through air. An off-
gas venting filter and a media filtration unit is supplied as part of the
treatment system.
The system treats flows of 26-45 L/min.
Ozomax: http://www.ozomax.com/Residential-products/point-of-use-ozone-systems.html
Ozone is produced electrolitically in situ using a precious metal
anode and an electricity supply. The manufacturers state that
higher ozone levels are formed at higher conductivity levels (up
to 600 mg/L total dissolved solids), so this method will not be
effective for soft water sources. The units come in a range of
sizes and can be fitted to domestic taps under the sink and also
incorporate a pre carbon filter to remove chlorine and a post
filter to remove precipitated solids. Monthly cleaning of the
system is required.
Veripure: http://www.veripurefood.com/index.php
The system is an in-line, low power ozone generation system (30
Watts) with a replaceable filter cartridge after ozonation to
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remove particulates from the water. The system is able to treat a flow of 1.9-15 L/min. No
information is provided as to how the ozone is generated.
Ozotech point of use ozone generator: http://www.ozotech.com/index.php/residential
A range of different sized systems are available for residential blocks
or under the sink application. Ozone is generated by a corona arc
discharge. Under the sink systems treat flows of 2.6 L/min, but the
residential system (POE) has a much greater capacity. Ozone is dosed
to around 1.5 mg/L, leaving a residual of 0.1-0.4 mg/L. The residential
system has an activated carbon post filter, whilst nothing seems to be
supplied with the under the sink ozonation systems. As with most
POU/POE devices, no detail on off gas destruction is mentioned.
DWC: http://www.dwc-water.com/technologies/ozone-disinfection/index.html
This company has a range of POU devices, but the one of most
interest is a system that generates ozone from water pressure
and therefore requires no electricity. The system is applied
directly onto the tap. The suppliers claim that it is able to
operate for 5000 hours, can treat 5 L per minute, supplying
ozone at a concentration of 0.2-0.3 mg/L.
POU devices are widely available from a range of suppliers (particularly in North America). Unless
installed immediately before water exits the tap, POU systems need to include an in-line filter to be
able to remove precipitated solids that will form as a result of the ozone application: not all of the
ozone generators reviewed have this capability. Off-gas destruction is also not considered in most
POU systems. This may be because all gas is transferred into the water or because accumulation of
off gas is low, but given that these systems are likely to be installed in customer’s homes this aspect
needs to be understood thoroughly. Further, appropriate independent testing of POU ozone devices
is required because to date, most, if not all, claims made by manufacturers have not been verified.
4.0 Application of ozone in wastewater treatment Similarly to drinking water treatment, ozone can be applied to satisfy a number of objectives in
wastewater treatment (examples of ozone applications in wastewater treatment are presented in
Table 10). Ozone is beginning to be employed in wastewater treatment for disinfection, oxidation,
improved membrane fluxes, reduced colour and odour, and removal of pharmaceutical and personal
care products (PPCPs) or other refractory organics (Sharif et al., 2012). Full scale ozone treatment
plants (with an ozone generation capacity of >0.5 kg/h) can be found being used for various
applications and treating almost all types of wastewaters:
Disinfection
Oxidation of inorganic compounds
Oxidation of organic compounds
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Enhancement of sludge degradability
Cost implications of full scale ozonation has led to the development of multistage treatment systems
as it has been shown that savings can be made by combining ozonation with biodegradation systems
(Gottschalk et al., 2000). Full scale wastewater ozonation systems usually employ bubble column
reactors as contactors and many are operated at elevated pressure (2-6 barabs)in order to achieve
high ozone transfer rates, which in turn increases the process efficiency (Gottschalk et al., 2000).
4.1 Disinfection of final effluent
Disinfection of waste waters is required in some countries, such as the US, before effluent is
discharged into receiving waters to meet water-quality standards. While bacteriological standards
are not currently part of the EU Urban Wastewater Treatment Directive, disinfection of municipal
effluents is implemented where there is a need to protect bathing waters in line with the EU Bathing
Water Directive. Chlorine is most commonly used but can cause formation of chlorinated
disinfection by-products (DBPs) (Gottschalk et al, 2010).
Although ozone disinfection is a well-established technology for drinking water treatment, ozone
disinfection mechanisms in wastewater are less well understood. However, where increasing
pressure on water resources has driven demand for water reuse for irrigation, ozone has been
utilised. For example, WWTP effluent in Almeria (Spain) is disinfected with ozone, before the water
is then applied to 3,189 ha of horticultural greenhouse crops. This provides water and nutrient
savings and eliminates pollutants from the wastewater discharge (Martinez 2011). Ozone reacts
strongly with many substances, therefore it is generally deemed more appropriate for use on pre-
treated effluents (Paraskeva and Graham, 2002).
A number of studies on ozone disinfection have utilised seeding approaches whereby laboratory
cultures of microorganisms (bacterial pathogens, viruses or indicators) are applied to buffer
solutions or wastewater which is then subject to ozonation. Reductions in faecal coliforms and
enteric viruses have been reported to undergo 3-6 log reduction in just a few seconds under these
conditions (Finch and Fairbairn, 1991: Herbold et al., 1989; Vaughn et al., 1987). For example,
Norwalk virus can be attenuated by ozone to a similar degree to the MS2 phage and was reported to
be removed to the order of 3-3.5 log units by 0.37-mg/L ozone for 10 seconds in ozone demand free
water (Shin et al, 2003). However, this may not give a realistic approximation of indigenous
organisms present in real wastewater treatment systems.
In wastewater systems, the ozone dosage required is dependent on the nature of the wastewater, as
for drinking waters (Gottschalk et al., 2010); therefore data from studies on actual wastewaters
provide a clearer indication of disinfection requirements. In laboratory semi-batch experiments,
carried out in a column reactor, ozone was applied to secondary treated effluent from a sewage
treatment plant in Varanasi, India (Tripathi et al., 2011). Following a conventional activated sludge
process, effluent was subject to a range of ozone doses and contact times at ambient temperature
(not given). Total and faecal coliforms and E. coli were inactivated at 10 mg/L at a 5 minute contact
time (Appendix 2) in effluent with a COD of 78 mg/L; turbidity of 40 NTU, pH 7.1 and characterised
as high in organic content. The authors noted that the applied ozone dose, rather than the
transferred dose was reported. Marked resistance of C. perfringens to ozone disinfection has been
noted by other authors (Tyrrel et al.1995). Ozone doses commonly presented in the literature
Page | 41
ranged from 0.3 µg/L to fully saturated, with contact times generally between 1.5 and 18 minutes.
This provided a range of inactivation rates, broadly in the range of 1-2.5 log inactivation for
coliforms, Enterococci and Clostridia, while some higher reductions were noted for bacteriophages
used as surrogates for human viruses (Appendix 2).
Xu et al. (2002) reported that hydraulic retention time had no significant effect on disinfection of E.
coli and faecal coliforms, rather than the critical parameter was optimisation of mass transfer of
ozone. Mass transfer tends to be low due to the low solubility of ozone (Zhou and Smith, 2000). For
example, Silva et al. (2010) reported transferred ozone percentages to wastewater of between 65
and 79 % in a bench-scale batch system. Many full-scale ozone reactors are operated at elevated
pressure (2–6 bar) to allow a high ozone-transfer efficiency, which also increases the overall process
efficiency (Gottschalk et al., 2010). When the transferred ozone dose approximates the immediate
ozone demand, before there is any measurable residual ozone, a significant log inactivation of faecal
coliforms and E. coli has been reported (Janex et al., 2000; Xu et al., 2002). Xu et al (2002) noted a 1-
3 log inactivation occurred when sufficient ozone was applied to meet the initial ozone demand.
Higher initial ozone demand can therefore mean higher inactivation rates at these doses, due to the
fact that the bacteria present form part of that ozone demand due to their high reactivity with
ozone. Xu et al. (2002) note that because of this, the CT approach used to calculate ozone
application for drinking water is inappropriate for disinfection of final effluent. At one of the same
wastewater treatment plants (in Indiannapolis) ozone disinfection had been successful, however no
ozone residual could be measured (Rakness et al. 2005).
For wastewater, ozonation is often part of a multistage treatment process to reduce ozone
consumption. Most often a chemical-biological process includes biodegradation at least before and
also often after the chemical oxidation step (Gottschalk et al, 2010). Ozone quantities and
generation costs tend to be reduced when utilising these combined systems.
Evaluating the combined effects of ozone and hydrogen peroxide, Gerrity et al (2011) established a
pilot HipOX reactor utilising and ozone/ H2O2 process for treatment of ultrafiltration (UF) or sand-
filtered secondary effluent after activated sludge treatment at a wastewater reclamation facility in
Nevada. Ozone was dosed at 5 mg/L and H2O2 at 3.5 mg/L with a residence time of 5 minutes. This
represented an ozone: total organic carbon ratio of 0.8. The addition of H2O2 promotes very rapid
decomposition of ozone into hydroxyl radicals, leading to completion of disinfection reactions within
seconds. Disinfection was evaluated for the sand filtration treatment train as UF was more effective
in eliminating coliforms. The ozone/ H2O2 process achieved 2-3-log inactivation of total coliforms and
3-4-log inactivation of faecal coliforms during two sampling campaigns. However, the ozone/H2O2
conditions were insufficient to comply with the effluent total coliform levels ranged from 284 to
1069 MPN per 100 mL. Faecal coliform levels ranged from 3.1 to 37 MPN per 100 mL, not meeting
US requirements for reuse. Limited inactivation Bacillus spores was reported (<1-log), however,
bacteriophage MS2 (human virus surrogate) underwent >6.5-log inactivation under the same
conditions. The authors stated that significantly higher ozone doses would have to be used to
achieve the 2.2 MPN per 100 mL threshold in the sand filtration configuration. However, UF pre-
treatment was effective in meeting this target. In contrast, Wert et al. (2007), comparing coliform
the efficacy of O3 and O3/H2O2 on tertiary treated wastewater reported that when similar O3
dosages are applied, O3 was a more effective disinfectant. They reported that sufficient disinfection
Page | 42
was achieved with O3 when applying dosages at or above the instantaneous ozone demand (IOD).
They recommended this for process control during full-scale application.
Subsequent BAC processing could also lead to regrowth of coliforms; therefore a final disinfection
process would potentially be required (Gerrity et al., 2011). Oh et al (2007) noted suggested that a
combined ozone/UV process could overcome the limitations of the ozone alone for the disinfection
of sewage effluent water.
While the generation of a measurable ozone residual for disinfection purposes poses problems, it
has been shown that combining ozone with a chlorine residual could be beneficial. Wert et al. (2007)
observed that ozonation reduced the formation of trihalomethane and haloacetic acids during
chlorination by 20%.
4.2 Impact of Ozone on filamentous bacteria
Reduction of excess sludge is a significant challenge for biological wastewater treatment and a good
understanding of the microbial ecology of the system is required to maintain an appropriate number
of filaments (Guo et al., 2012). Filamentous organisms are a normal part of the activated sludge
microflora and, in low numbers, are thought to promote floc formation. Excessively long filaments or
presence in high numbers can lead to sludge bulking (Eckenfelder, 1992).
Ozone can be used as a “non-specific” approach to reducing filamentous organisms and therefore
reducing bulking and foaming in activated sludge systems. Chlorine is commonly used for this
purpose, but doesn’t represent an on-going solution as it can affect nitrification (Thirion, 1982).
Partial ozonation of the return sludge of an activated sludge process can also substantially reduce
the excess sludge. During ozonation, in addition to damaging and inactivating microorganisms
present in sludge, organic matter is converted into biodegradable materials and there is some
mineralization of soluble organic matter through chemical oxidation (Foladori et al, 2010). Ozonation
also improves settling properties of the sludge and reduces bulking and foaming (Vergine et al,
2007). Ozonation has been applied directly within biological wastewater treatment processes or as
pre-treatment prior to anaerobic digestion (Yazoo and Shibata, 1994; Weemaes et al., 2000; Goel et
al., 2004; Lee et al., 2005; Dytczak et al., 2007) in order to reduce excess sludge. Low dose ozonation
can inhibit the activity of filamentous bacteria and has been applied to control bulking and improve
floc settling (Foladori et al, 2010, Appendix 3a).
However, there is relatively little understanding of the mechanisms involved and, as with any
approach adopted for reducing sludge production, it influences the microbial community which
could affect effluent quality (Vergine et al, 2007). Organisms responsible for bulking and foaming are
listed in Table 18 below.
The efficiency of sludge ozonation depends on the following parameters (Foladori et al; 2010):
Wastewater or Sludge quality
Reactor configuration
Ozone gas flow rate and concentration
Sludge flow rate and solids concentration
Ozone transfer efficiency
Page | 43
Contact time
Ozone dosage per mass of TSS
Table 18. Filamentous bacteria found in bulking activated sludge (adapted from Guo and Zhang, 2012).
Types Phylum Genus or species according to the reference Reference related to bulking
Beggiatoa Proteobacteria Beggiatoa Williams and Unz, 1985
Cyanobacteria Cyanobacteria Unknown N.A.
Flexibacter elegans CFB group Flexibacter elegans Jenkins et al., 1993
Haliscomenobacter hydrossis Bacteroidetes H. hydrossis Ziegler et al., 1982
Leucothrix mucor Proteobacteria Leucothrix mucor Williams and Unz, 1989
Microthrix cali Actinobacteria Candidatus Microthrixa Levantesi et al., 2006
Microthrix parvicella Actinobacteria Candidatus Microthrixa Rossetti et al., 1997
Mycobacterium fortuitum Actinobacteria Mycobacterium fortuitum Blackall et al., 1991
Nocardiaform-like organisms Actinobacteria Gordonia amarae Blackall, 1994
Nostocoida limicola Firmicutes Trichococcus Liu et al., 2000
Nostocoida limicola Actinobacteria Tetrasphaera jenkinsii Liu and Seviour, 2001
Nostocoida limicola Planctomycetes Isosphaera Liu et al., 2001
Rhodococcus globerulusc Actinobacteria Rhodococcus globerulus Schuppler et al., 1995
Rhodococcus ruberc Actinobacteria Rhodococcus ruber Schuppler et al., 1998
Skermania piniformisc Actinobacteria Skermania piniformis Eales et al., 2006
Sphaerotilus natans Proteobacteria Sphaerotilus Eikelboom, 1975
Thiothrix form I Proteobacteria Thiothrix Howarth et al., 1999
Tsukamurella pseudospumaec Actinobacteria T. pseudospumae Nam et al., 2004
Type 0041/0675 TM7-likeBetaproteobacteria UnknownAquaspirillum
Hugenholtz et al., 2001Thomsen et al., 2002; Thomsen et al., 2006
Type 0092 Chloroflexi CFB Unknown Speirs et al., 2009Bradford et al., 1996
Type 0211 Unknown Unknown N.A.
Type 021N Proteobacteria Thiothrix eikelboomii Howarth et al., 1999
Type 0411 CFB group Unknown Bradford et al., 1996
Type 0581 Unknown Unknown N.A.
Type 0803 Chloroflexi Caldilinea Kragelund et al., 2011
Type 0914 Chloroflexi Unknown Speirs et al., 2011
Type 0961 Unknown Unknown N.A.
Type 1701b Betaproteobacteria Unknown Howarth et al., 1998
Type 1851 Kouleothrix aurantiaca Chloroflexi (Phylum) Beer et al., 2002
Type 1863 ProteobacteriaCFB group AcinetobacterMoraxella osloensisChryseobacterium Seviour et al., 1997
It is apparent that low dose ozonation has a differential effect on filamentous bacteria (Caravelli et
al, 2006), hence its ability to reduce the organisms responsible for filamentous bulking without
affecting biological treatment (van Leeuwen and Pretorius, 1988; Vergine et al, 2007). It is likely that
filaments protruding from flocs come into immediate and direct contact with ozone, whereas those
Page | 44
within the floc are protected from the immediate effects, therefore ozone-damage occurs more
substantially in the filamentous organisms (Seviour and Neilsen, 2010). Chu et al. (2009a) observed
that there was no immediate decrease in ATP upon exposure to ozone, but immediate decrease in
enzyme activities. This indicates that ozone first affects the filaments, destroying the floc, then
disperses compact aggregates (potentially affecting enzymes embedded in EPS or on cell surfaces)
and then affects bacterial cells leading to the decrease in ATP (Chu et al, 2009a). Foladori et al.
(2010) noted that at high dose ozonation tended to reduce the number of small flocs (< 3 µm) and
increase the number of medium flocs (7.5-30 µm) but that particle size generally remained stable at
lower doses. ESEM images have shown a deformation of bacterial cells subjected to <0.17 g O3/g TSS
(Chu et al., 2008).
Yan et al. (2009) reported that there was little evidence of any impact on microbial communities,
based on molecular analyses, for ozone dosages up to 20 mg O3/g. Chu et al. (2009) noted that a
systematic study of the changes in characteristics and activities of sludge exposed to low-dose ozone
would be useful for optimization of a cost-effective sludge reduction process. In earlier studies, van
Leeuwen (1987), using microscopic methods, noted differences in the types of filamentous
organisms present, with apparently greater diversity in the ozonated sludge. The number of
filaments appeared to be an order of magnitude lower in the ozonated treatment. The authors also
indicted that ozonation promoted nitrification and biological removal of organic material without
affecting phosphate removal.
Figure 4. Floc density with and without treatment (from Wijnbladh, 2007).
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Figure 5. Filament presence with and without treatment (from Wijnbladh, 2007).
Wijnbladh (2007) observed that ozonation of activated sludge led to changes in the microscopic
appearance of the filamentous bacterium M. parvicella in the reactor. Initially, the organism stained
Gram positive and exhibited substantial storage of phosphate granules. However, after three weeks
of ozonation of the return activated sludge, the bacteria appeared either Gram positive with many
empty cells or even completely Gram negative. No changes were noted in a parallel control
treatment train. The authors reported improved floc structure (Figure 4) directly after start up, with
few M. parvicella remaining inside the sludge flocs (Figure 5). Similar observations were made two
months from start-up of treatment. No effects on nitrification were observed in the study. Likewise,
Paul and Debellefontaine (2007) reported that ozone caused filamentous bacteria to disappear,
leading to the reduction of bulking and foaming in activated sludge and generating more compact
flocs.
The influence of ozone on other wastewater treatment parameters is critical to the success of its use
to reduce filaments or excess sludge. Some reported impacts on other microorganisms or
biologically driven processes are given in Appendix 3b. Paul and Debellefontaine (2007) noted that
there was a linear relationship between the log of biomass activity (reported as maximal oxygen
uptake rate) and log ozone dose between 0.001 and 0.2 g O3 transferred per g COD in the sludge).
They concluded that ozonation does not affect any of the capabilities of an AS biological process;
however other studies have reported contrary indications.
A decrease in nitrification rate has been shown to be proportional to increasing ozone dose (Dytczak
et al. (2007); Appendix 3b). The literature is contradictory in the degree to which nitrification is
affected. For example, van Leeuwen (1988) and Vergine (2007) did not experience any reductions in
nitrification or total nitrogen removal at ozone doses of 1-4 mg/g MLSS/d and 6 mg/L applied to
settled sewage or activated sludge respectively. In contrast, Böhler and Siegrist (2004) observed a
reduction in the rate of nitrification capacity similar to the reduction in sludge production across
Page | 46
doses between 0.02 and 0.08 g O3/g TSS and Kobayashi reported an 80 % reduction in nitrification
rate at a rate of 0.05 g O3/g TSS. It has been suggested that the decrease in nitrification, due in part
to direct effects of ozone on nitrifiers and in part to competitive interactions with heterotrophs, may
be offset by the increased sludge retention time occurring as a result of the decrease in excess
sludge production (Paul and Debellefontaine, 2007; Foladori et al., 2010). Indeed, Déléris et al (2002)
observed similar ammonia removal in both ozonated (0.062 g O3/g TSS) and control activated sludge
treatments.
The viability of phosphorus accumulating organisms (PAOs) and heterotrophic bacteria has been
reported to decrease exponentially with the degree of sludge solubilization during ozonation in the
range of 0.03–0.04 g O3/g TSS (Saktaywin et al., 2005). This was corroborated by Lee et al (2005) who
reported that an ozone dose of 0.05 g O3/g TSS, inactivated 97 % of heterotrophic organisms and
Yasui and Shibata (1994) who observed a 90 % decrease in colony forming units at less than 0.05 g
O3/g TSS. Some novel approaches to biological nutrient removal also have the potential to combine
ozonation with other technologies to provide effective wastewater treatment while minimising
excess sludge production. For example, Suzuki et al. (2006) proposed an anaerobic-oxic-anoxic
system equipped with an ozonation tank and a P adsorption column. Trials demonstrated that
despite a slight decrease in nutrient removal efficiencies, 34-127 % sludge reduction was obtained
and 80% of P was recovered.
Reid et al (2007) suggested that a more cost effective ozone process for excess sludge reduction
would follow the principle “partial oxidation as low as possible and biological oxidation as high as
possible”. This minimises the use of ozone but maximises conversion of solids into biodegradable
materials which can then be removed in cheaper biological reactors.
Cao et al. (2009) set out to investigate the effect of ozone on the formation of typical disinfection by-
products during application of chlorine residual to tertiary treated wastewater for reclamation. This
is also pertinent where treated effluent significantly influences raw water for abstraction. Ozonation
was carried out in batch reactors prior to application of a chlorine dose of 5 mg/L. Ozonation led to
an increase in the formation of all the DBPs measured, including THMs and HAAs, except CHCl3,
which decreased from 51.2 to 37.1 µg/L at consumed ozone doses of 0-10 mg/L. Total brominated
THMs increased from 10.3 to 58.2, primarily through increases in CHCl2Br and CHClBr2. This was
thought to be due to the relatively high concentration of organic compounds in the tertiary
wastewater. In contrast, Wert et al. (2007) found that ozonation reduced the formation of THMs
and HAAs during chlorination by 20%. The same authors also reported that when applying an
advanced oxidation processes with H2O2 and ozone, the lack of O3 residual coupled with the
presence of excess H2O2 inhibited bromate formation. Furthermore, it has been reported that an
advanced oxidation process such as combined O3-UV has the potential to remove DBP precursors
from raw surface waters. Chin et al (2005) concluded that this combined AOP was more effective
than either technology alone and reduced THM formation by around 80 % and HAA formation by
around 70 % at 0.62 mg O3/ML and 1.61 Ws/cm2 UV dosage. Ozonation has also been reported to
reduce the genotoxicity (based on the umu assay which relates to DNA damage) of the treated
wastewater, both with and without chlorination. Genotoxicity decreased from 7 µg 4-NQO/L with
neither chlorine nor ozonation, to just over1 µg 4-NQO/L with 10 mg/L ozone and 10 mg/L chlorine
Page | 47
(Cao et al, 2009). Takanashi et al. (2002) also noted that ozone treatment was effective in removing
mutagen precursors from wastewater.
It is clear that the nature of the water to be treated is critical in determining the degrees of DPB
formation.
4.3 Oxidation of inorganic compounds
One of the aims of ozone application in wastewater treatment is to remove toxic inorganic
substances and this mainly involves the removal of cyanide (CN-) mostly associated with metal
processing and electronics industry wastewaters. It can be present in its free form which reacts
quickly with ozone but more often it is complexed with iron or copper (Gottschalk et al., 2000). The
complexed forms of cyanide are more stable to ozone attack and therefore advanced oxidation
processes are more appropriate for its removal.
Nitrite (NO2-) and sulphide (H2S/S2-) react quickly with ozone and therefore their removal is
sometimes carried out using ozonation, however more cost-efficient biological treatment
alternatives are more often employed for these contaminants.
4.4 Oxidation of organic compounds
Most often in industrial wastewater treatment, ozone is applied to remove target organic
compounds that can be present at a wide range of concentrations. These wastewaters include
landfill leachate, textile, pharmaceutical and chemical industry wastewaters that can contain many
refractory organics including humic compounds, aromatic compounds containing metals, pesticides
and surfactants. The main aims of ozonation in this case are (Gottschalk et al., 2000):
The transformation of toxic organics that are often present in low concentrations and as
complex mixtures
The improvement of biodegradability of refractory organics by partial oxidation
The removal of colour
Usually complete mineralisation by ozone is not economical and combination with other processes is
preferable. The most frequent problems associated with wastewater ozonation include foaming and
precipitation of inorganic salts such as calcium oxalate, calcium carbonate and ferrous hydroxide
which can lead to reactor clogging. A number of full-scale ozonation systems has been used in
Germany to treat landfill leachate prior to effluent discharge to water bodies and due to the
complex nature of the organics the ozonation stage is commonly operated between two biological
systems (Bio-O3-bio) (Gottschalk et al., 2000). The main aim of ozone application in the treatment of
textile wastewaters is removing non-biodegradable colour but additionally the removal of
surfactants and partial oxidation of DOC can be achieved but again a multistage treatment system is
usually necessary. Low ozone doses are sufficient for the removal of colour but if a high degree of
DOC removal is required the treatment costs increase dramatically.
A number of studies have focused on the coupling of ozone and aluminosilicates to remove volatile
organic carbon compounds (VOCs) (Brodu et al., 2012). In this case methyl ethyl keton was used in
the treatment studies and showed that ozonation regeneration is a promising step in VOC removal
(Brodu et al., 2012).
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The use of ozone in water reuse applications is becoming increasingly popular due to its efficacy in
removing trace organics with limited to no chemical addition and the potential for reduced energy
requirements (Gerrity and Snyder, 2011). Ozone and OH radicals are extremely effective in
oxidizing most trace organic compounds, particularly the steroid hormones linked to feminisation of
aquatic species, but the required level of treatment in each application is relatively subjective
(Gerrity and Snyder, 2011). The potential of ozonation as a tertiary treatment prior to industrial
water re-use (paper mill effluent, textile wastewater, food and chemical industry effluents) was
recently investigated at lab scale with promising results (Mauchauffee et al., 2012). Several issues
are associated with the use of ozone in wastewater including the formation of bromate, N-
nitrosodimethylamine (NDMA), and other potentially toxic oxidation by-products (Gerrity and
Snyder, 2011). Some utilities are mitigating the potential effects of ozone by-products through
subsequent biological filtration with acclimated sand or activated carbon (Gerrity and Snyder, 2011).
Reliability and maintenance issues has hampered the application of ozone in wastewater treatment
plants and the limited experience makes it difficult to extrapolate existing design knowledge, which
is primarily based on drinking water applications, to wastewater matrices with their higher oxidant
demand, turbidity, organic content, UV absorbance, and radical-scavenging capacity (Gerrity and
Snyder, 2011).
Due to low ozone reactivity with a number of compounds, advanced oxidation processes that
promote the formation of OH has been studied in recent years. One such process is photocatalytic
ozonation (TiO2/UV/O3) which has been studied to treat inorganic anions, such as cyanide ions,
simple organic molecules, such as monochloroacetic acid, oxalate anions and formic acid, and
environmentally problematic organic compounds, such as pesticides, additives, and textile dyes (Sun
et al., 2013). Bobu et al. (2013) employed ozone based AOPs (O3/UV, O3/UV/H2O2 and
O3/UV/H2O2/Fe(II)) for the degradation of two antibiotics (enrofloxacin and ciprofloxacin)achieving
the highest degradation with the O3/UV/H2O2/Fe(II) treatment (93 and 99% TOC removal
respectively).
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Table 19. Ozone applications in wastewater treatment.
Reference WW/substrate treated
System scale
Treatment mode
Ozone dose applied
Contact time Removal Comments
Gago-Ferrero et al., 2013
Benzophenone-3 (BP3)-personal care product in DI water
Lab-scale Batch 60 min >99% Initial conc. (BP3)= 5.1 mg/L; BP3 shows higher reactivity through radical pathways compared to direct ozonation
Li et al., 2007 Benzophenone-3 (BP3) in tertiary effluent
Full-scale Continuous 5-6 mg/L 180 min 20% Initial conc. (BP3)= 311 ng/L
Jagadevan et al., 2013
Spent metalworking fluids (model wastewater)
Lab-scale Batch 2500 mg/L 27% after ozonation; 72% overall
Initial COD 3100 mg/L; ozonation followed by biological stage
Bertanza et al., 2011
Bisphenol A in WW Pilot-scale 8 mg/L 80 min 90% Initial conc. BPA = 2 - 4.3 x 10-4 mg/L
Mohapatra et al., 2012
Bisphenol A in WW sludge
Lab-scale Batch 24.14 mg/g SS
16.47 min 100% 24 g/L SS, pH 6.23
Muz et al., 2013 EDCs (diltiazem, carbamazepine, butyl benzyl phthalate, acetaminophen, estrone and progesterone) in activated sludge
Lab-scale Batch 1.1 mg O3/L 4 x 6 min >99% Ozone pulsing of waste sludge on 4 successive days reduced aerobic digestion from 30 to 4 days with MLSS reduction over 80% in the same period
Qiang at al., 2012 EDCs (estrone (E1), estriol (E3), 17a-ethynylestradiol (EE2), bisphenol A (BPA), and 4-nonylphenol (NP)) in activated sludge
Lab-scale Semi-batch 100 mg O3/g SS
majority of estrogens removed; BPA 83%; NP 64%
E1, EE2, E3 and BPA spiked at 2 g L-1; NP conc. in activated sludge
adequately high (ca. 76 g/L)
Turhan and Ozturkcan, 2013
anionic sulphonated azo dye (Reactive orange 16) in synthetic WW
Lab-scale Semi-batch 24 mg/L 8 min Complete best treatment efficiency achieved at basic conditions (pH 12)
Qian et al., 2013 Bio-treated textile wastewater
Lab-scale Batch 3.1 mg O3/mg COD
5 min Turbidity 95.8%, colour 97.5%, COD 88.1%, DOC 68.7% UV254 90.5%
Combined treatment process (coagulation + O3/GAC) studied here
Page | 50
Table-continued……Ozone applications in wastewater treatment Reference WW/substrate
treated System scale
Treatment mode
Ozone dose applied
Contact time Removal Comments
Gagnon et al., 2008
Primary effluent containing trace organics (salicylic acid, clofibric acid, ibuprofen, 2-hydroxy-ibuprofen, naproxen, triclosan, carbamazepine, and diclofenac)
Pilot-scale 20 mg/L 18 min > 70% apart for ibuprofen and 2- hydroxyl ibuprofen
The primary effluent: pH 8.1 - 8.2, TSS 5 mg/L, DOC 90 - 110 mg/L, and residual Al (0.6 - 0.9 mg/L) and Fe (0.3 - 0.4 mg/L)
Wert et al., 2009 Tertiary treated WW containing range of trace organics
Pilot-scale Continuous O3/TOC = 1 24 min > 80%
Sundaram et al., 2009
Secondary effluent with 30 monitored trace organics
Pilot scale continuous 7 mg/L All > 90% High bromate formation reported
Hollender et al., 2009
WW effluent containing 220 monitored trace organics
Full-scale continuous O3/DOC = 0.6 4-10 min Only 11 compounds detected > 100 ng/L-1
Page | 51
4.5 Enhancement of sludge degradation
Ozonation is a widely used chemical method to improve anaerobic degradability of sludge. However,
sludge ozonation was first used in combination with the activated sludge process for wastewater
treatment and a treatment system based on these processes has been commercialised by the
Japanese Kurita company (ca. 30 installations have been implemented) and another industrial
process has been proposed by Ondeo-Degremont (Suez): the Biolysis® O process (Carrère et al.,
2010). Ozonation has also been combined with anaerobic digestion as a pre-treatment or post-
treatment with a recycle back to the anaerobic digester. Better performance and lower ozone
consumption has been observed in the case of post-treatment and recycling in the digester (Carrère
et al., 2010). Disintegration of sludge by ozonation as a pre-treatment process to accelerate the
digestion processes has been of recent interest (Erden et al., 2010). An ozone dose of 0.1 g O3/kg
total solids increased the aerobic degradability of the sludge with the volatile suspended solids (VSS)
reduction achieved in the digester increasing to 34% in comparison to 22% reduction without the
ozonation pre-treatment. Ozonation pre-treatment did not have any major effect on the
dewaterability of the sludge. The advantages of ozonation pre-treatment of secondary sludge is that
it can improve the sludge solubilisation and it can also simultaneously degrade organic pollutants
such as bisphenol A (Mohapatra et al., 2012).
5.0 Environmental Implications It is thought that the changes brought about through ozonation of wastewater may modulate or
possibly increase the toxicity of effluents toward aquatic organisms. The availability and mobilisation
of heavy metals and other chemicals could be affected by ozonation, notably where dissolved
organic matter contains substantial amounts of protein or humic acids. Direct effects of ozonation
on shrimp hatching rates have been observed (Sellars et al., 2005). However, it is the indirect effects
of altered wastewater chemistry that are most likely to pose a risk to aquatic biota. For example, T
lymphocyte proliferation in Rainbow Trout decreased dramatically in fish exposed to ozonated
effluent compared to fish exposed to either the primary-treated effluent or to aquarium water
(Hebert et al., 2008).
However, at the appropriate dosage, beneficial effects of ozonation have been highlighted. For
example, Wei et al (2012) took samples from a wastewater treatment plant and a connected
reclamation treatment plant in Beijing to evaluate some of the treatment processes from a
toxicological view point. Ozonation was applied to secondary effluent at 0, 5, 10 and 15 mg/L with
contact times of 0, 5, 10, 15 and 20 minutes, after which residual ozone was removed. Further to
this, ozonated effluent (contact times up to 45 minutes) was passed through an activated carbon
filter. Effluent from both ozone only and ozone plus activated carbon were pre-treated and
biotoxicity tests applied. For 15 mg/L of ozone dose and 5 min contact time, acute toxicity and
genotoxicity decreased the former to a greater extent. Combining ozone application with activated
carbon treatment provided more effective reduction in toxicity.
More in-depth research is required to better understand interactions between ozone and
wastewaters and their impacts on the ecology of receiving waters.
Page | 52
6.0 Conclusions & Recommendations Ozone disinfection of drinking water is fairly well established across arrange of countries in
Europe and within the US. It has been demonstrated to be highly effective against a wide
range of pathogens, including the particularly resistant organism, Cryptosporidium. There is
potential for the use of ozone in combination with a chlorine residual to protect against
regrowth within the distribution network and for regulatory monitoring where detection of
a residual is required.
A key consideration of ozone treatment for any of the above processes is the ability to
deliver the appropriate dosage efficiently. This is one of the major challenges with ozonation
due to its low solubility. Improved ozone transfer technologies will enhance the cost
effectiveness of this process. Deriving the appropriate dosage for drinking waters
disinfection is likely to be straightforward through the utilisation of a range of available
models and provided attention is given to influent water quality data. In contrast, it is likely
to be more difficult to define dosing for wastewater applications, where the standard CT
approaches do not hold.
A further key consideration is the relationship between ozone and water quality. Because
ozone is such a powerful oxidant, it is important to evaluate the outcome of its application
to waters at a given point in the treatment train and to determine whether this placement a)
makes the most of the effect of ozone in terms of both disinfection and degradation of
pollutants, and b) will not result in the formation and subsequent discharge of unacceptable
levels of disinfection by-products. Or, if this is likely to occur, whether it is possible and /or
cost-effective to add on a further treatment step to deal with those additional pollutants.
Implementation of ozone treatments is likely to be most effective if evaluated on a site by
site basis as this not only promotes a thorough evaluation of how to implement the
technology for effective disinfection, but would also allow the operators to determine
whether ozone interventions at multiple points would be more cost effective for overall
treatment and/or whether the inclusion of additional treatment stages before or after
ozonation would be required.
Cyst and sporular forms of protozoans and bacteria present the most difficult treatment
challenge for ozone disinfection. High CT values can control these pathogens, but it is
recommended that a physical barrier is also provided for water sources containing these
micro-organisms.
The main limitation of applying ozone in drinking water systems is the production of
bromate as a DBP. Methods of controlling bromate formation have been developed,
principally through controlling pH and understanding prevailing water quality conditions.
Empirical models are one way by which bromate formation can be predicted and controlled,
but these models are only accurate when developed for a specific water source.
POU/POE systems are widely available, mainly from North American suppliers, that can
generate and deliver ozone from mains electricity. A filtration system needs to be supplied
after the ozone when these are used in order to prevent precipitated solids from being
present in treated water – it is not always evident that the proprietary systems have these as
supplied. Who manages and replaces spent filtration systems and adequate off-gas control
must therefore be also considered for POU systems.
Ozonation of final effluent can also be effective, however higher ozone doses tend to be
required which not only increases cost but also increases the likelihood of formation of
disinfection by products. Careful integration with other treatment processes such as
Page | 53
biological filtration can allow this approach to work, however and Potential to couple
biological and filtration based treatments with ozone treatment and disinfection towards a
more cost-effective approach to use of ozone at WWTPs.
Ozone has been shown to be an effective non-specific inhibitor of filamentous organisms in
activated sludge systems and has provided effective long term management of bulking and
foaming sludge. While there is a limited amount of literature relating to this application,
most reports are positive in terms of effectiveness in restoring floc structure and minimising
filaments. There is on-going debate of the effect on other microbiota within the system and
therefore of effects upon nitrification and phosphate accumulation. Further research would
help to elucidate the parameters controlling the range of process effects.
A wide range of organic and inorganic contaminants can be degraded by ozone. High
concentrations of ozone are needed for effective degradation of bulk natural organic matter
and is therefore not recommended for this application. The most effective use of ozone is
for oxidation of metals and organic micropollutants in combination with a
physical/adsorptive process.
The higher contaminant and scavenging load in wastewater means that much higher doses
must be applied for these waters in order to achieve satisfactory levels of
removal/disinfection. There are many examples of ozone having been used in wastewater
for small scale and pilot treatment systems, but few cases of large WWTWs using ozone due
to the high cost of having to add such high concentrations of ozone.
Ozone is a widely used chemical in water and wastewater treatment as well as numerous
industrial applications. The ability of ozone to effectively oxidise a wide range of
contaminants and disinfect a broad sweep of micro-organisms have made it an essential
component of many treatment flowsheets across the world.
7.0 Future Research and Direction Understanding water quality interactions with ozone is critical to its effective use for inactivation of
any unwanted microorganisms across the spectrum of drinking, wastewater and reclamation
treatment systems. Existing water quality data for influent waters and effluents from each
treatment stage should provide valuable information within which to evaluate where within a given
system to implement ozonation to best and most efficient effect. The development of a simple
decision tree based on predominant water quality trends could be utilised to fulfil this purpose.
Further trials on realistic pilot scale systems to develop process based rules to relate water
parameters to all round treatment efficacy provide a useful and realistic research direction,
alongside continuing development of ozone contactor systems to provide the highest efficiency
transfer.
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8.0 References Amy, G., Siddiqui, M., Zhai, W. and Debroux, J. (1994) Survey on bromide in drinking water and impacts on DBP formation. Denver, Colorado, American Water Works Association (Report No. 90662).
Audenaert, W.T.M., Callewaert, M., Nopens, I., Cromphout, J., Vanhoucke, R., Dumoulin, A., Dejans, P., Van Hulle, S.W.H. (2010) Full-scale modelling of an ozone reactor for drinking water treatment Chemical Engineering Journal, 157 (2-3), 551-557.
Beer, M., Seviour, E.M., Kong, Y., Cunningham, M., Blackall, L.L. and Seviour, R.J. (2002). Phylogeny of the filamentous bacterium Eikelboom Type 1851, and design and application of a 16S rRND targeted oligonucleotide probe for its fluorescence in situ identification in activated sludge. FEMS Microbiology Letters, 207, 179-183.
Beltrán, F.J., García-Araya, J.F. and Álvarez, P.M. (1999) Integration of continuous biological and chemical (ozone) treatment of domesticwastewater, 2. Ozonation followed by biological oxidation. Journal of Chemical Technology and Biotechnology, 74, 884–890.
Bertanza, G., Pedrazzani, R., Dal Grande, M., Papa, M., Zambarda, V., Montani, C., Steimberg, N., Mazzoleni, G., Di Lorenzo, D. (2011) Effect of biological and chemical oxidation on the removal of estrogenic compounds (NP and BPA) from wastewater, an integrated assessment procedure. Water Research 45, 2473–2484.
Blackall, L.L. (1994) Molecular-identification of activated-sludge foaming bacteria. Water Science and Technology, 29, 35-42.
Blackall, L.L., Harbers, A.E., Greenfield, P.F. and Hayward, A.C. (1991) Foaming in activated-sludge plants -a survey in Queensland, Australia and an evaluation of some control strategies. Water Research, 25, 313–317.
Blanc, D., Carrara, P., Zanetti, G. and Francioli, P. (2005) Water disinfection with ozone, copper and silver ions, and temperature increase to control Legionella, seven years of experience in a University teaching hospital. Journal of Hospital Infection, 60, 69-72.
Bobu, M., Yediler, A., Siminiceanu, I., Zhang, F., & Schulte-Hostede, S. (2013) Comparison of different advanced oxidation processes for the degradation of two fluoroquinolone antibiotics in aqueous solutions. Journal of Environmental Science and Health. Part A, Toxic/Hazardous Substances & Environmental Engineering, 48(3), 251–262.
Böhler, M. and Siegrist, H. (2004) Partial ozonation of activated sludge to reduce excess sludge, improve denitrification and control scumming and bulking. Water Science and Technology, 49, 41–49.
Bose, P and Reckhow, D.A. (2007) The effect of ozonation on natural organic matter removal by alum coagulation. Water Research, 41, 1516-1524.
Bouland, S., Duguet, J.-P., Montiel, A. (2004) Minimizing bromate concentration by controlling the ozone reaction time in a full-scale plant. Ozone: Science and Engineering 26 (4), 381-388.
Bradford, D., Hugenholtz, P., Seviour, E.M., Cunningham, M.A. , Stratton, H. Seviour, R.J., Blackall, L.L. 16S rRNA analysis of isolates obtained from gram-negative, filamentous bacteria micromanipulated from activated sludge. Systematic and Applied Microbiology, 19. 334-343.
Brodu, N., Zaitan, H., Manero, M.-H., & Pic, J.-S. (2012) Removal of volatile organic compounds by heterogeneous ozonation on microporous synthetic alumina silicate. Water Science and Technology : a Journal of the International Association on Water Pollution Research, 66(9), 2020–2026.
Broséus, R., Barbeau, B. Bouchard, C. (2008). Verification of Full-Scale Ozone Contactor Inactivation Performance Using Biodosimetry. Journal of Environmental Engineering, 134, 304-315.
Buffle, M.-O., Schumacher, J., Meylan, S., Jekel, M. and von Gunten, U. Ozonation and advanced oxidation of wastewater: effect of O3 Dose, pH, DOM and HO scavengers on ozone decomposition and HO generation. Ozone Science and Engineering, 28, 247-259.
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Burkhari, Z., Marshall, M.M., Korich, D.G., Fricker, C.R., Smith, H.V., Rosen, J. and Clancy, J.L. (2000). Comparison of Cryptosporidium parvum Viability and Infectivity Assays following Ozone Treatment of Oocysts. Applied and Environmental Microbiology, 66, 4598.
Butler, R., Godley, A., Lytton, L. and Cartmell, E. (2005) Bromate environmental contamination: review of impact and possible treatment. Critical Reviews in Environmental Science and Technology, 35, 193-217.
Campbell, I.S., Tzipori, S., Hutchinson, G. and Angus, K.W. (1982). Effect of disinfectants on survival of Cryptosporidium oocysts. Veterinary Record, 111, 414-415.
CaO, N., Miao, T., Li, K., Zhang, Y. and Yang, M. (2009). Formation potentials of typical disinfection byproducts and changes of genotoxicity for chlorinated tertiary effluent pretreated by ozone. Journal of Environmental Sciences, 21, 409-413.
Caravelli, A., Giannuzzi, L., and Zaritzky, N. (2006). Effect of ozone on filamentous bulking in a laboratory scale activated sludge reactor using respirometry and INT dehydrogenase activity. Journal of Environmental Engineering, 132, 1001–1010.
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Appendix 1: A synthesis of Ozone inactivation of Cryptosporidium in Water and Wastewater Process Ozone
(mg/L) contact time (min)
CT value (mg.min/ L)
Inactivation Rate (%)
Conditions Additional notes Reference
Primary disinfection - - 40
25-35 <25
99.9 >90<99 <90
Laboratory batch reactor Buffer soln. pH7; 1 ºC
Bovine source Oocysts pH effects. Old Oocysts may have been more susceptible to Ozone
Driedger. (2001)
- - 22 15-18 <15
99.9 >90<99 <90
Laboratory batch reactor Buffer soln. pH7; 5 ºC
- - 12 8-10 <8
99.9 >90<99 <90
Laboratory batch reactor Buffer soln. pH7; 10 ºC
- - 1.5 1 <1
99.9 >90<99 <90
Laboratory batch reactor Buffer soln. pH7; 20 ºC
Sequential disinfection with ozone and free chlorine
- - 1.4 90 Laboratory batch reactor Buffer soln. Greater synergy was observed at pH 6 than at pH 7.5, and no synergy was observed at pH 8.5.
Bovine source Oocysts pH effects. Old Oocysts may have been more susceptible to Ozone
Inactivation in demand free water
1T 3 5 10
3c 5-10 10c
90 90-99 99-99.9
Laboratory experiments. Ozone demand-free water at pH7; 25ºC.
Oocysts from calves.Based on mouse infectivity assays and excystation assays
Korich et al. (1990).
Inactivation in DI water
4 T 7 21c ~99 Laboratory study in DI water; 20ºC; pH 7. Oocysts from calf. Based on fluorescent staining and SEM.
Ran et al. (2010) 2 T < 10 <20c ~99
Inactivation in DI water of differing turbidity (Kaolinite added)
3 T 7 21c 99.2 Laboratory study in DI water Turbidity 0.1 NTU; 22ºC; pH 7
Inactivation in DI water of differing turbidity (Kaolinite added)
3 T 7 21c 86.2 Laboratory study in DI water Turbidity 20 NTU; 22ºC; pH 7
Inactivation in DI water, differing pH
3 T 5 15c ~ 99 Laboratory study in DI water; 20ºC;pH 6
Inactivation in DI water, differing pH
3 T 5 15c ~ 39 Laboratory study in DI water; 20ºC; pH 9
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Inactivation in DI water, differing DOM
3 T 7 21c 99.1 Laboratory study in DI water; 20ºC; pH7; Turbidity 1 NTU; DOM 0 mg L-1.
Inactivation in DI water, differing DOM
3 T 7 21c 98.3 Laboratory study in DI water; 20ºC; pH7; Turbidity 1 NTU; DOM 1 mg L-1.
Inactivation in DI water, differing DOM
3 T 7 21c 84.9 Laboratory study in DI water; 20ºC; pH7; Turbidity 1 NTU; DOM 2 mg L-1.
Inactivation in DI water, differing DOM
3 T 7 21c 62.1 Laboratory study in DI water; 20ºC; pH7; Turbidity 1 NTU; DOM 10 mg L-1.
Natural river water - 7.4* 6 99 24.5 ±1.6 ºC; pH 8.24 ±0.20 Owens et al. (2000) - - 3.5 99 22ºC; pH 6.9 Finch et al., (1993) - - 4.02-4.62 99 22ºC; pH 7 Peeters et al., (1989) Batch liquid, modified batch ozone
0.6 4 2.4-3.2 99 25 ºC Langlais et al., (1990)
Semi Batch system - - 5.39 99 20.8 ºC, pH=7, Phosphate buffer Rennecker et al., 1999 (cited by Facile et al. ,2000)
Advanced water treatment plant. Continuous ozone generation in contact tank. Water extracted and added to seeded containers.
2.5 (~0.3R)
5 7.5c 66 24 ºC; pH 7.1 No information on whether filtration was before or after ozonation. No water quality data.
Wohlsen et al., 2007
2.5 (~0.05R)
10 25c 92 24 ºC; pH 7.1
2.5 (~0.7R)
5 7.5c 82 18.9 ºC; pH 6.9
2.5 (~0.15R)
10 25c 93 18.9 ºC; pH 6.9
Laboratory experiment; 80 ml stirred beakers.
0.3 2 0.6 c 1-85
22 9 ºC; pH; 6.3-6.7 Ozonated DI water;107 oocysts per beaker. Comparison of methods for detecting inactivation of oocysts. Up to 99% differences in outcome dependent on method used.
Bukhari et al (2000).
0.4 2 0.8c 0-99 22 9 ºC; pH; 6.3-6.7
Clarified and clarified filtered secondary municipal effluent
15 10 150c < 15 17-27 (mean 20 and 22) ºC; pH 6.7-8.6. Primary screening, sedimentation,pre ppt. with pAlCl3; secondary – AS and sedimentation; clarification (experimental samples removed here), flocculation with 30-40mg L-1 pAlCl3; 6 h sedimentation, chlorine.
Liberti et al. (2000)
Natural waters - - 52 99 3 ºC; pH 6.2-8.2; turbidity 0.2-2.5. Experimental and modelled data. Range of conditions.
Oppenheimer et al (2000) - - 29 99 7 ºC; pH 6.2-8.2; turbidity 0.2-2.5. - - 18 99 10ºC; pH 6.2-8.2; turbidity 0.2-2.5. - - 3.9 99 20ºC; pH 6.2-8.2; turbidity 0.2-2.5. - - 2.9 99 22ºC; pH 6.2-8.2; turbidity 0.2-2.5. - - 1.8 99 25ºC; pH 6.2-8.2; turbidity 0.2-2.5.
*Theoretical contact time T = transferred ozone concentration; R = residual ozone concentration C= calculated by Avery as product of prior two columns. Ozone dose is given as applied dose unless otherwise stated.
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Appendix 2: A synthesis of Ozone disinfection of final effluent
Process Ozone (mg/L)
contact time (min)
CT value (mg.min/ L)
Organism Inactivation Rate Conditions / noted Reference
Wastewater Disinfection (WWTP)
5 T 1.5 7.5c MS2 Coliphage 6.5 log removal HiPOXTM Reactor Microfiltred water
Ishida et al (2008)
- - 1 Total coliforms Not detected
Natural Water subject to inputs of sewage and agricultural drainage
Sat 5-10 - Aerobic mesophiles >2-log At higher starting densities ~ 108 CFU ml-1 Aerobic mesophiles; 103-105 CFU ml-1 Total Coliforms Spore forms appeared to disappear after 15 minutes but re-emerged after 30 minutes indicating germinationhad occurredgerminate
Voidaru et al (2007)
Sat 30 - Aerobic mesophiles 7-8-log (not detected)
Sat 5-10 - Total coliforms >2-log
Sat 15-20 - Total coliforms 3.5-4-log (not detected)
Sat 10-15 - Faecal coliforms >2-log
Sat 15 - Faecal coliforms 2.5 log (not detected)
Sat 10-15 - Enterococcus >2-log
Sat 15 - Enterococcus 2.5 log (not detected)
Sat 15 - C. perfringens (cells) 1 log (not detected) Secondary sewage Effluent (after activated sludge)
0.3R ~2 0.6c Faecal coliforms 1.4 Single pulse Tyrrell et al (1995)
0.3 R ~2 0.6c Enterococci 1.1
0.3 R ~2 0.6c C perfringens 0.1
0.3 R ~2 0.6c F+ coliphage >2.4
0.3 R ~2 0.6c Somatic coliphage >1.9
0.3 R ~2 0.6c Faecal coliforms 1.1
0.3 R ~2 0.6c Enterococci 1.0
0.3 R ~2 0.6c C perfringens 0.2
0.3 R ~2 0.6c F+ coliphage >2.8
0.3 R ~2 0.6c Somatic coliphage 2.2 Secondary sewage Effluent (after rotating biological contactor)
0.3 R ~2 0.6c Faecal coliforms 1.5
0.3 R ~2 0.6c Enterococci 1.2
0.3 R ~2 0.6c C perfringens 0.1
0.3 R ~2 0.6c F+ coliphage >2.8
0.3 R ~2 0.6c Somatic coliphage >2.8 Secondary sewage Effluent (after activated sludge)
0.3 R ~2 0.6c Faecal coliforms 1.5
0.3 R ~2 0.6c Enterococci 1.2
0.3 R ~2 0.6c C perfringens 0.1
0.3 R ~2 0.6c F+ coliphage >2.2
0.3 R ~2 0.6c Somatic coliphage >2.1 Secondary effluent after activated sludge
5 5-10 25-50 c Total coliforms, Faecal coliforms, E. coli 1.1-1.4 COD 78 mg L-1 40 NTU, pH 7.1 “high” organic content. Ambient temperature (India).
Tripathi et al (2011)
10 5-10 50-100 c Total coliforms, Faecal coliforms, E. coli 2.5-2.6
15 5-10 75-150 c Total coliforms, Faecal coliforms, E. coli 2.8-3 Anaerobic effluent from an Upflow Anaerobic Sludge
5 (3.8T) 5 25 c Total coliforms E.coli
2.2 2.6
Silva et al (2010)
5 (4.0 T) 10 50 c Total coliforms 2.2
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Blanket (UASB) reactor E.coli 2.7 5 (4.0 T) 15 75 c Total coliforms
E.coli 2.4 2.9
8 (5.4 T) 5 40 c Total coliforms E.coli
3.0 3.2
8 (5.2 T) 10 80 c Total coliforms E.coli
3.0 3.2
8 (5.7 T) 15 120 c Total coliforms E.coli
3.0 3.5
10 (7.1
T) 5 50 c Total coliforms
E.coli 3.2 4.0
10 (7.1
T) 10 100 c Total coliforms
E.coli 3.6 4.2
10 (7.7
T) 15 150 c Total coliforms
E.coli 3.7 4.3
Tertiary wastewater effluent
4.9 T 18 88 Total coliforms Faecal coliforms
>3 >3
Wert et al. (2007)
7.3 T 18 131 Total coliforms Faecal coliforms
>3 >3
8.7 T 18 157 Total coliforms Faecal coliforms
>3 >3
2.1 T 6 13 Total coliforms Faecal coliforms
>1 >2
3.6 T 18 65 Total coliforms Faecal coliforms
> 3 > 3
7.1 T 10 71 Total coliforms Faecal coliforms
>2 > 3
Ozone dose is assumed to be applied dose unless denoted T to indicate transferred or consumed ozone. Sat = saturation. C denotes contact time calculated using dose x time.
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Appendix 3a: Ozone impacts on filamentous bacteria
Process Ozone Dose
contact time (min)
Organism Inactivation Rate or other effect
Conditions Additional notes Reference
AS AS
18 µg g-1 VSS
5-30 Whole floc respiration reduction
54-60 % Caravelli et al. (2006)
18 µg g-1 VSS
5-30 Filamentous bacteria respiration reduction
87 % Filamentous bulking was controlled under these conditions.
AS 1-4 g- kg-MLSS-1 d-1
- Sludge volume index decrease
50 mL g-1
Continuous dosing No effect on nitrification-denitrification even at30 mg g-1 MLSS d-1
Van Leeuwen (1987)
AS for biological N removal
0.05-0.1 kg kg TS-1 (T)
- Excess sludge reduction Microthrix parvicella reduction Nocardioforms reduction
39% 3-5 to <2 (Jenkins scale) 1-3 to 0-1(Jenkins scale)
Wastwater dominated by textile waste Anoxic pre-denitrification basin; aerobic nitrification. Praxair system applied to part of RAS stream. Ozonated sludge returned to nitrification basin.
No effect on COD and TN removal Disappearance of biological foaming
Vergine et al (2007)
Sudge from municipal WWTP in Beijing
20 mg L-1 15 Oxygen uptake rate Decrease
59% Bubble column, batch operation, lab scale. TSS 4400-4800 mg L-1; 63-72 % VSS.
Continuous ozonation Chu et al (2009a)
Settled sewage, pilot plant, Municipal WWTP, Pretoria
3 mg L-1 (T) - Filamentous organisms reduced SS Reduction
Little effect Little effect
Continuous dosing Van Leeuwen and Pretorius, (1988)
Settled sewage, pilot plant, Municipal WWTP, Pretoria
6 mg L-1 (T) - Filamentous organisms reduced SS Reduction
1 order of magnitude 33 %
Continuous dosing Increased settleability of sludge;no effects on N or P removal, increased organics removal
Van Leeuwen and Pretorius, (1988)
Excess sludge production
0.01 g O3 g TSS-1
- Excess Sludge production reduced.
50% Intermittent ozonation of AS Required only 30% of the ozone dose required for continuous ozonation.
Kamiya and Hirotsuji (1998)
0.02 g O3 g TSS-1
- Excess Sludge Reduction
100%
Ozone dose is assumed to be applied dose unless denoted (T) to indicate transferred or consumed ozone.
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Appendix 3b: Ozone impacts on other microbiota or microbial processes
Process Ozone (g ozone/g TSS unless otherwise stated)
Inactivation Rate Reference
Phosphate accumulating organisms 0.03-0.04(T) Exponential decrease Saktaywin et al., 2005 Heterotrophic bacteria 0.03-0.04(T) Exponential decrease Heterotrophic bacteria 0.05 97 % Lee et al (2005) Heterotrophic bacteria <0.05 90 % Yasui and Shibata (1994) Nitrifying bacteria 0.05 80 % Kobayashi et al., 2001 (cited by Chu et al 2009b) Pathogens - coliforms >0.1 Significant effectiveness Foladori et al (2010) Streptococcus and Salmonella 0.2-0.4 100% Park et al (2008) Electron transport activity 0.04 Increased slightly and reduced quickly Zhao et al. (2007) Dehydrogenase activity >0.04 May be affected Nishimura et al., 1999 Maximum O2 uptake rate 0.9-13.6 mg O3
g COD(T) Decreased Dziurla et al. (2005) Alteration in bacterial community DNA 0.02(T) No effect Yan et al. (2009) 0.03-0.06(T) Some species lost >0.06(T) Only two strains survived 0.08(T) DNA failed to amplify (inactivated)
Ozone dose is assumed to be applied dose unless denoted (T) to indicate transferred or consumed ozone.
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Appendix 4: Disinfection Byproduct Formation during Ozonation of water and wastewater
Process Water Characteristics (mean or approximate) Ozone treatment DBPs formed Reference pH Alkal-
iInity COD TOC TSS TS NH4
+ BrO3- Br- Nitrate temper
ature Dose (mg/L) Time DBP (ug/L)
USAB Effluent 6.5 183 82 388 5-10 5-15 Aldehydes upto 187 Silva et al. (2010) 6.5 183 82 388 5-10 5-15 Glyoxal Upto 46 Tertiary Wastewater
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 <3.6 Upto 24 Bromate <2 Wert et al. (2007)
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 3.6 -<5 Upto 24 Bromate <10
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 5-7.1 Upto 24 Bromate >20<40
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 >8.5<11.5 Upto 24 Bromate 50-65
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 2.1 6 Aldehydes Carboxylic acids
0 396
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 3.6 18 Aldehydes Carboxylic acids
114 623
7.0 120 - 7 - - < 0.08
<0.001
0.25 14 20 7.0 18 Aldehydes Carboxylic acids
211 397
Tertiary Wastewater – ozonated then chlorinated (5 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 0(T) - CHCl3 CHCl2Br CHClBr2 CHBr MCAA MBAA DCAA DBAA TCAA
51 5 3 1 2 1 14 2 9
Cao et al (2009)
Tertiary Wastewater – ozonated then chlorinated (5 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 1 (T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
47 18 14 1 2 1 22 2.5 13
Tertiary Wastewater – ozonated then
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 3 (T) - ChCl3 ChCl2Br ChClBr2
38 16 15
Page | 73
chlorinated (5 mg L-1)
ChBr MCAA MBAA DCAA DBAA TCAA
2 3 1 21 3 8
Tertiary Wastewater – ozonated then chlorinated (5 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 5 (T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
29 15 20 3 3 2 26 4 12
Tertiary Wastewater – ozonated then chlorinated (5 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 10 (T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
28 22 34 4 4 2 32 5 14
Tertiary Wastewater – ozonated then chlorinated (10 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 0(T) - CHCl3 CHCl2Br CHClBr2 CHBr MCAA MBAA DCAA DBAA TCAA
84 19 18 2 4 1 42 1 28
Tertiary Wastewater – ozonated then chlorinated (10 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 1 (T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
80 38 20 1 5 2 49 2 34
Tertiary Wastewater – ozonated then
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 3(T) - ChCl3 ChCl2Br ChClBr2
80 41 30
Page | 74
chlorinated (10 mg L-1)
ChBr MCAA MBAA DCAA DBAA TCAA
2 5 2 57 3 35
Tertiary Wastewater – ozonated then chlorinated (10 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 5 (T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
78 40 32 2 5 2 55 4 36
Tertiary Wastewater – ozonated then chlorinated (10 mg L-1)
7.2 - 9.1 6.7 (DOC)
5 - 0.6 0.7
- - 10(T) - ChCl3 ChCl2Br ChClBr2 ChBr MCAA MBAA DCAA DBAA TCAA
50 39 31 2 6 2 53 5 27
Ozone dose is assumed to be applied dose unless denoted (T) to indicate transferred or consumed ozone.
CREW Facilitation Team
James Hutton Institute
Craigiebuckler
Aberdeen AB15 8QH
Scotland UK
Tel: +44 (0) 844 928 5428
Email: [email protected]
www.crew.ac.uk